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Abstract

Waste Water Treatment Plants (WWTPs) are receptors for the cumulative loading of microplastics (MPs) derived from industry, landfill, domestic waste water and storm water. The partitioning of MPs through the settlement processes of waste water treatment results in the majority becoming entrained in the sewage sludge. This study characterised MPs in sludge samples from seven WWTPs in Ireland which use anaerobic digestion (AD), thermal drying (TD), or lime stabilisation (LS) treatment processes. Abundances ranged from 4,196 to 15,385 particles kg-1 (dry weight). Results of a general linear mixed model (GLMM) showed significantly higher abundances of MPs in smaller size classes in the LS samples, suggesting that the treatment process of LS sheer MP particles. In contrast, lower abundances of MPs found in the AD samples suggests that this process may reduce MP abundances. Surface morphologies examined using Scanning Electron Microscopy (SEM) showed characteristics of melting and blistering of TD MPs and shredding and flaking of LS MPs. This study highlights the potential for sewage sludge treatment processes to increase or reduce the risk of MP pollution prior to land spreading and may have implications for legislation governing the application of biosolids to agricultural land.
Microplastics pathway through sludge treatment processes
338x190mm (96 x 96 DPI)
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Microplastics in Sewage Sludge: Effects of Treatment
1
A.M. Mahon
a
*, B. O’Connell
a
, M.G. Healy
b
, I. O’Connor
a
., R. Officer
a
, R. Nash
a
, L.
2
Morrison
c
3
a
Marine and Freshwater Research Centre (MFRC), Galway-Mayo Institute of Technology, Dublin Road,
4
Galway, Ireland.
5
b
Civil Engineering, National University of Ireland, Galway, Ireland.
6
c
Earth and Ocean Sciences, Schools of Natural Sciences and Ryan Institute, National University of Ireland,
7
Galway, Ireland.
8
*Corresponding Author: Anne Marie Mahon: annemarie.mahon@gmit.ie 9
10
Abstract
11
Waste Water Treatment Plants (WWTPs) are receptors for the cumulative loading of
12
microplastics (MPs) derived from industry, landfill, domestic waste water and storm water.
13
The partitioning of MPs through the settlement processes of waste water treatment results in
14
the majority becoming entrained in the sewage sludge. This study characterised MPs in
15
sludge samples from seven WWTPs in Ireland which use anaerobic digestion (AD), thermal
16
drying (TD), or lime stabilisation (LS) treatment processes. Abundances ranged from 4,196
17
to 15,385 particles kg
-1
(dry weight). Results of a general linear mixed model (GLMM)
18
showed significantly higher abundances of MPs in smaller size classes in the LS samples,
19
suggesting that the treatment process of LS shear MP particles. In contrast, lower abundances
20
of MPs found in the AD samples suggests that this process may reduce MP abundances.
21
Surface morphologies examined using Scanning Electron Microscopy (SEM) showed
22
characteristics of melting and blistering of TD MPs and shredding and flaking of LS MPs.
23
This study highlights the potential for sewage sludge treatment processes to affect the risk of
24
MP pollution prior to land spreading and may have implications for legislation governing the
25
application of biosolids to agricultural land.
26
27
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Keywords: Microplastics; sewage sludge; biosolids; anaerobic digestion; lime stabilisation;
28
thermal drying.
29
30
1. Introduction
31
Microplastics (MPs) are synthetic polymers measuring less than 5 mm in diameter and are
32
derived from a wide range of sources including synthetic fibres from clothing,
1,2
polymer
33
manufacturing and processing industries,
3
and personal care products.
4
They have the
34
potential to adsorb persistent organic contaminants
5,6
and priority metals
7-11
from the
35
surrounding environment. These may be released upon digestion by biota or through
36
environmental degradation, leading to possible impacts to human health and ecosystems.
12-14
37
Over the last 10 years, many studies have investigated the distribution
1,15
and effects
16-19
of
38
MPs within the marine environment. Indeed, MPs have been found in Polar Regions
20
and in
39
a range of freshwater environments worldwide.
21-24
Despite this, few studies have sought to
40
determine land-based sources of MPs.
25
Wastewater treatment plants (WWTPs) have been
41
identified as receptors of MP pollution and effective in capturing the majority of MPs in the
42
sludge during settlement regimes
26
, as first found by Habib et al. when they used synthetic
43
fibers as a proxies for the presence of sewage
27
. More than 10 million tonnes of sewage
44
sludge was produced in WWTPs in the European Union (EU) in 2010.
28
European Union
45
policy on sustainability and recycling of resources
29
favours the recycling of sludge. The
46
introduction of EU legislation such as the Landfill Directive (1999/31/EEC
30
) and the
47
Renewable Energy Directive (2009/28/EC
31
) have diverted sewage sludge from landfill and
48
incineration into use for energy production
32
and agriculture.
33
In some countries, such as
49
Ireland, up to 80% of municipal wastewater sludge is reused in agriculture.
34,35
Guidelines
50
stipulate that the sludge must undergo some type of treatment (after which it is commonly
51
referred to as ‘biosolids’) prior to land application. This may include lime stabilisation (LS),
52
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anaerobic digestion (AD), composting, or thermal drying (TD).
31
As approximately 99% of
53
MPs are retained in sewage sludge generated in WWTPs,
36
there is a possibility that land
54
applied sludge, even having undergone treatment, could be a source of MPs pollution.
55
56
The regulations for the use of biosolids in the EU and USA stipulate limit levels for pathogen
57
content, maximum metal and nutrient application rates to land
37
and vector (flies and rodents)
58
attraction reduction (USA only). Restrictions in land application of biosolids vary between
59
the EU and USA. Under US federal legislation, the application of biosolids to agricultural
60
land can occur without restriction in volume or duration, if the contamination level reaches an
61
exceptional quality “EQ”.
37
The concentration limits vary for all contaminants and to date
62
MP pollution is not included.
37
In Europe, sewage sludge is dealt with very differently
63
among member states, and application to land is banned in some countries.
38-40
64
65
As most sewage sludge undergoes treatment prior to land-spreading, the effects of these
66
treatments on microplastic morphology is important but remains largely unknown, with some
67
evidence of increased abundance of fibres at a smaller size range for LS sludge
41
which was
68
attributed to alkaline hydrolysis, facilitated by elevated pH, heat and mechanical mixing.
41,42
69
Therefore, the aim of this study was to investigate the first stage of the MP pathway post-
70
WWTP, and the impacts of different treatments. In particular, it aimed to determine if (1)
71
MPs are present in treated sewage sludge from a range of WWTPs employing AD, TD and
72
LS as treatment techniques, and (2) the type of treatment used (TD, AD, LS) employed at the
73
WWTP impacts on MP abundance and characteristics, including size and surface
74
morphology.
75
2. Methodology
76
77
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2.1 WWTP sludge sample collection and preparation
78
Sewage sludge, having undergone treatment including TD, AD or LS, was collected from
79
seven waste WWTPs with population equivalents (PEs) ranging from 6500 to 2.4 million
80
(Table 1). These WWTPs received waste water from industry, storm water run-off and
81
domestic sources, all of which comprised up to 30% of the influent organic loading
82
(measured as biochemical oxygen demand, BOD) (Table1). Due to an intense difficulty in
83
gaining access to these WWTPs and the primary focus being in the chacteristion of sewage
84
sludge applied to land (post-treatment), pre-treatment samples were not taken. Three replicate
85
samples of 30 g were obtained from each WWTP and stored at -20
o
C prior to sample
86
preparation. The treated sewage sludge had dry matter (DM) contents ranging from 24%
87
(AD) to 87% (TD). Pellets of TD sludge were placed in water for 1 week to induce softening,
88
transferred to a water bath (30
o
C) for 24 hr, and placed in an “end-over-end” shaker
89
(Parvalux, UK) for 12 hr. This shaking procedure was repeated until the pellets were
90
sufficiently softened without compromising the physical characteristics of the MPs. The
91
samples were subsequently washed through a 250 µm sieve, which resulted in complete
92
degradation of the pelleted clumps prior to elutriation. As handling and identification of MPs
93
becomes unfeasible below 200 µm using available instrumentation, and most studies have
94
examined MPs above this size,
44
the sieve size (250 µm) was selected to optimise efficiency
95
of MP extraction and identification. Although fibres are only 10-20 µm in diameter, they are
96
generally longer than 250 µm and curved to some degree and the majority are therefore
97
trapped in the 250 µm sieve.
20
A proportion of the washed through fraction (WTF) was
98
retained and passed through 212, 63, and 45 µm sieves for particle size determination or
99
particle size fractionation. These sieve sizes were chosen to include the majority of smallest
100
particles and therefore give an impression of treatment effects of sludge particle size.
101
102
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Anaerobically digested and LS sludge were soaked in filtered tap water to soften and
103
homogenise them, and were also washed through 250, 212, 63 and 45 µm sieves to determine
104
particle size fractions. As the LS sludge had an oily appearance, thought to be derived from
105
the break-down of cellulosic material through alkaline hydrolysis, it was decided that the
106
elutriation and other density separation techniques were unsuitable for extraction of MPs.
107
Instead, 10 g from each replicate sample were examined by passing it directly through a filter
108
(GF/C: Whatman
TM
, 1.2 µm) using vacuum filtration.
109
110
2.2. Microplastics Extraction
111
2.2.1 Elutriation
112
The principal of elutriation was used as the first step in the separation of MPs from other
113
sample components. Elutriation separates lighter particles from heavier ones through an
114
upward flow of liquid and/or gas, and has been widely used in the separation of biota within
115
sediment samples.
42
To separate MPs from the sludge samples, an elutriation column, based
116
on the design of Claessens et al.
43
was constructed.
117
118
2.2.1.1 Column extraction efficiency estimation
119
To check for efficiency of the column in extracting MP, three sediment samples, each
120
weighing 40 g, were spiked with 50 microplastic particles of high density polyethylene
121
(HDPE) (three colours) and PVC, and run through the column. The HDPE samples used were
122
shavings of approximately 1.0 (L) × 4.0 (W) × 2.0 mm (B). The PVC particles were of a
123
similar dimension, but were more brittle. Therefore, each particle was marked with a blue
124
marker to ensure that particles were not counted twice upon recovery. The number of
125
particles, separated from the sediment matrix, that exited the column, was enumerated and the
126
percentage efficiency was calculated.
127
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128
2.2.2. Zinc chloride (ZnCl
2
) extraction
129
The microplastic extraction was filtered through 250 µm mesh, rinsed into a separatory
130
funnel with 1 molar ZnCl
2
solution, and brought to a volume of 300 ml. The funnel was
131
plugged, vigorously shaken for 1 min, and allowed to settle (20 min). The settled material
132
was drained and the remainder of the sample was filtered onto glass fibre filters (GF/C:
133
Whatman
TM
, 1.2 µm). The oily appearance of the LS samples rendered this density
134
separation technique unsuitable for extraction of MP.
135
136
2.3. Characterisation of MPs
137
The filters were examined using stereomicroscopy equipped with a polariser (Olympus
138
SZX10) attachment and a Qimaging
®
Retiga™ 2000R digital camera. Microplastics were
139
identified and enumerated based on several criteria including form, colour and sheen used in
140
previous studies as described by Hidalgo-Ruz et al.
44
For fibers, the form of a synthetic fibre
141
should not taper at either end, while not having rigidly straight form. Any polymer will not
142
have cellular structure or other organic structures. Artificial fibres/ particle also have
143
uniformity of colour and exhibit a sheen once pass through the polarized light. Where
144
ambiguity remained following these observations, the suspected polymer was manipulated
145
with a hot pin by which a melted form indicated a positive result. Microplastics were
146
measured and allotted to the following size categories: 250-400 µm, 400-600 µm, 600-1000
147
µm, and 1000-4000 µm. Suspected microplastics were enumerated and measured and
148
approximately 10% of MP specimens from each filter paper were set aside for polymer
149
identification. This was deemed sufficient as the cumulative abundance of MP concentration
150
within the sludge systems made it unfeasible to examine each MP particle. As MPs were
151
randomly selected, they covered a range of size-classes and morphological types.
152
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Microplastics for which any ambiguity remained as to whether it was a polymer or not were
153
automatically selected for analyses.
154
155
Attenuated total reflectance (ATR) and Fourier transform infrared spectroscopy (FTIR)
156
(Perkin Elmer, USA, Spectrum Two
TM
with Universal ATR Accessory and Thermo
157
Scientific, UK, Nicolet iN10 FTIR microscope with germanium Tip Slide-on-ATR) was used
158
to analyse approximately 10% of MP specimens. The spectra were obtained with 3-second
159
data collection (16 scans per sample) over the wave number range 600 – 4000 cm
-1
using a
160
liquid nitrogen-cooled MCT-A detector at 8 cm
-1
resolution. SEM analyses was carried out in
161
order to examine the surface structures of MPs of LS and TD sludges. A selection of MP
162
samples (~20) extracted from the sludge (and pristine plastics for comparative purposes) were
163
gold-coated (Emitecg K550, Quorum technologies, Ltd., UK) and subjected to variable
164
pressure scanning electron microscopy (SEM) in secondary electron mode using a Hitachi
165
model S2600N (Hitachinaka, Japan). The analyses were performed at accelerating voltages of
166
10 - 20 kv; an emission current (I
c
) of 10 µA; working distance of 12 - 24mm.
45
MP samples
167
chosen where mostly those which were easily divided in 2 to allow for simultaneous polymer
168
identification and SEM analyses. There was therefore a bias to larger MPs
169
170
2.4 Quality control and contamination prevention
171
Cotton laboratory coats and nitrile gloves were used during the sample preparation and
172
analyses. In addition, synthetic clothing was avoided and samples were covered at all times
173
and working surfaces were cleaned with alcohol prior to use. When analysing filter papers, a
174
blank filter paper was exposed to the open laboratory conditions to assess the possibility of
175
air-borne contamination.
176
177
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2.5. Data analyses
178
Statistical analyses were carried out using Minitab 17 (2010) and R (R Core Team, 2012). As
179
data were not normally distributed, non-parametric tests were used to test for differences in
180
MP abundances amongst locations (Mann-Whitney Test). To investigate if there were any
181
possible effects of PE on abundance, a Spearman’s rank correlation analysis test was utilised.
182
With the exception of one WWTP, there was only one treatment method employed per site
183
(Table 1), so in-site correlation was not possible. Each site was treated as an independent
184
measurement and plotted using a box plot. A GLMM (generalised linear mixed effect model;
185
Eqn. 1) was used to investigate the high number of MP particles in the smaller class sizes at
186
WWTPs in which LS was used.
187
188
!"#$%&'()*"#+#%,-*) . /$0(*10-*+/2&0 3 4%&,'(*"%-+56,"7('0-* 3
8
9:;<=>;?=+@A<?=
+
189
Eqn. 1
190
191
Where 1/Treatment Plant specifies a random intercept model.
192
193
A separate GLMM for each size class was carried out using a Poisson distribution and a
194
random effect term to account for nesting of replicates within WWTPs to determine which
195
explanatory variable was responsible for larger proportions of smaller MP particles at lime
196
stabilised WWTPs.
197
198
199
200
201
3. Results and Discussion
202
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3.1 Characterisation of treated sewage sludge
203
The characteristics of the sewage sludge treated using AD, LS and TD had varying physical
204
characteristics. The particle size fractionation (g/kg) of the AD samples was smaller than the
205
LS and TD samples (Table 2), and had a sandy appearance. The AD samples were very dark
206
and heavy with some cellulosic material, whereas the TD samples had a lot of cellulosic
207
material entrained, which was difficult to separate during elutriation and zinc chloride
208
extraction. Although this cellulosic material was distinctive from MP material (in that its
209
fibres tapered at the ends and it was often branched) and therefore easy to disqualify, its
210
presence in the samples increased greatly the time and consumables (filter papers) utilised
211
during the filtration process. High levels of cellulose derived from toilet paper in sewage may
212
merit the inclusion of a digestion process using the cellulase enzyme as has been previously
213
used for the isolation of MPs in North Sea sediments.
46
214
215
216
3.2 Microplastics Extraction
217
3.2.1 Elutriation column extraction efficiency estimation
218
The average extraction efficiency rate of the elutriation column for the spiked sediment
219
samples was 90%, 94% and 91% for the red, blue and black HDPE particles, respectively.
220
The elutriation process was less efficient for the PVC particles, which resulted in an average
221
extraction efficiency of 80%. This is an indication that results of MP abundance in this study
222
may be an underestimation. As the efficiency test was carried only for fragments at one size
223
only, it may not be representative of efficiency of fibre removal.
224
225
3.3 Characterisation of MPs
226
3.3. 1 Microplastics abundance
227
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MPs extracted from the biosolids ranged from an average of 4,196 to 15,385 particles kg
-1
228
(DM) among the seven sites, with significant differences in MP abundances between some
229
sites and within Site 1 (1A, 1B) between AD samples and TD samples (Mann Whitney, w =
230
15, p = 0.0809; Figure 1). This is likely to be an underestimation due to loses in column
231
efficiency (approx. 20%) and through the use of a 250 µm sieve from which a proportion of
232
fibres may be lost.
20
Airbourne contamination on the control filter papers was negligible (1-2
233
fibres per day) due to high abundance of MPs in these samples. The use of a different
234
method using a smaller sample size after sieving (250 µm) for LS samples (sites 5,6), may
235
have had an effect on the quantification as no loss of MPs would have occurred during the
236
elutriation process. This may have led to the possible inflation of MP abundances in LS
237
samples compared to TD and AD sludges. The abundances found in this study are in the
238
same order of magnitude as the study by Zubris and Richards
40
which reported between 3,000
239
and 4,000 particles kg
-1
. In this study, a lack of correlation between PE and MP abundance
240
kg
-1
(Spearman’s rank, r = - 0.308, p = 0.458) implies that these differences may have been
241
due to the variation of source inputs (industrial, storm water, landfill etc.). However, as no
242
data exists on the temporal variation of MPs in sewage sludge, the possibility also exists that
243
these variations are a result of fluxes in MP input which could be a result of peak MP
244
emission times in relation to household and industrial activity etc. In this study, the
245
significantly lower abundance of MPs reported for an anaerobically digested bio-solid sample
246
(1B) which was collected from the same site as sample 1A (TD) compared to all other sample
247
except site 3 (also AD) (taken roughly at the same time), posits an interesting question over
248
the possible role of AD in the degradation of MPs. Without pre-treatment samples, there is no
249
evidence to prove that the Mesophilic anaerobic digestion (MAD) used at the AD treatment
250
plants in this study, facilitated the breakdown of MPs and few studies have look at the
251
breakdown of polymers in anaerobic digesters. However, one pilot study which investigated
252
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the effect of plastic waste on the functioning of anaerobic digestion, finding that digesters
253
from which plastic was removed produced less gas that those to which plastic was added.
47
254
As there is already substantial evidence of microbial breakdown of polymers through the
255
activity of exoenzymes (promoting depolymerisation) and assimilation of smaller articles
256
resulting in mineralisation,
48 49,50
, the role of degradation by microorganisms within the AD
257
systems should be further investigated.
258
259
3.3.2 Morphological categorization and polymer identification of MPs
260
This study confirmed that MPs are retained in the sewage sludge and are largely composed of
261
fibres, similar to what was found by Talvite et al.
47
and Magnusson and Norén.
36
262
Approximately 75.8% of the MP consisted of fibres, followed by fragments, films, other
263
unidentified particles, and spheres, which accounted for only 0.3% of total MP abundance
264
(Table 3). The greatest proportion of MP fragments was found at the LS WWTPs, with Site 6
265
being the only site to have marginally more fragments than fibres (Table 3; Figure 2).
266
Polymers, identified by FTIR, comprised HDPE, polyethylene (PE) polyester, acrylic,
267
polyethylene terephthalate (PET), polypropylene, and polyamide (Figure 3). Some of these
268
contained minerals. Although waste water derived from households generate high quantities
269
of fibres, principally derived from clothes washing of >1900 fibres per wash
1
, other industrial
270
sources of fibres such as the fibre manufacturing industry may also be important contributors.
271
272
273
274
3.3.3. Size of MPs
275
Using the fitted coefficients from the GLMM, hypotheses of no difference between all
276
pairwise combinations of the treatment effects were tested. At small and medium particle
277
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sizes, the LS treatment was significantly different from both TD and AD treatments (Figure
278
4; P < 0.001; sizes classes A and C; P < 0.05 size class B). During the analyses, raising the
279
MP characteristics (abundance and size), from the smaller sample size processed for LS
280
samples (10 g) to the equivalent of that in the TD and AD samples (30 g), may have reduced
281
the representativeness of the size-classes within LS samples. However, from the sludge
282
particle size fractionation, where 40 g were sampled from LS, TD and AD sludges, there was
283
also a larger proportion of sludge (g) in size classes <45 µm, > 45 µm and > 63 µm,
while
284
lower sludge mass at the larger size fraction >212 µm (3 g compared to 32 and 35
285
respectively). This corresponds to the higher number of MPs at smaller size classes in LS
286
samples derived from our analyses and therefore suggests that the mechanisms that resulted
287
in smaller sludge particle size may be responsible for MP size distribution alike. In this study,
288
as it was not possible to obtain pre-treatment samples, it is not possible to wholly assign the
289
differences in size classes to the treatment processes. However, the elevated numbers at the
290
small size classes for LS samples, are in agreement with results reported by Zubris and
291
Richards.
40
In their study, there was evidence of elevated abundance of MPs at smaller size
292
classes for LS samples which were derived from the same sludge pre-treatment batch. The
293
authors attributed this to elevated pH combined with mechanical mixing. In a more recent
294
study, Cole et al.
17
found partial destruction of nylon fibres and melding of polyethylene
295
fragments using 10M NaOH at 60
o
C. It is likely, therefore that in this study, the combination
296
of elevated pH, temperature and mechanical mixing could be responsible for the elevated
297
numbers of MPs in smaller sizes classes. As alkaline-stabilized sludges have been associated
298
with higher levels of metals compared to other sludges
51
, the possibility of shearing effects
299
should be investigated further to identify potential impacts.
300
301
3.3.4 Surface morphologies of MPs
302
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Scanning electron micrographs of surface textures of polymers entrained in the treated
303
biosolids had some surface morphologies, which varied among treatment type. An unknown
304
polymer fibre, which was thermally dried, had distinct blistering and fracturing, particularly
305
in the fibre curves (Figure 5: A-C). Additionally, polymer fragments from TD samples,
306
identified as HDPE and PE fragments, showed wrinkling, melding and some fracturing,
307
which was quite distinct from pre-treatment samples (Figure 6: G-I; Figure 7: D-F). Surface
308
morphologies of MPs originating from LS biosolids had a more shredded and flaked
309
appearance for the unknown polymer (Figure 5: D-F) and a HDPE specimen (Figure 5: D-F).
310
This is in accordance to some degree with the study by Zubris and Richards
40
, which reported
311
surface morphologies exhibiting high abrasion and appeared “very brittle”.
Anaerobically
312
digested samples of an unknown polymer had deep cleavage, which was distinct from any
313
other observations (Figure 5: G-I). It is not possible to wholly assign these observed surface
314
morphologies to effects treatment processes due to low replication and the absence of
315
comparison of surface morphologies of these MPs prior to entering the treatment plant.
316
However, these images provide some idea of initial fracturing patterns which can be observed
317
in MPs which have travelled thus far in their pollution pathway.
318
319
4. Conclusions
320
Although it was not possible to assign wholly the abundances or size distributions to the
321
treatment processes, results suggest that treatment processes may have an effect. If MPs are
322
altered by treatment, the potential for impact may also be influenced accordingly. This could
323
add to the unknown risks associated with MPs in sewage sludge. Regardless of treatment
324
regimes, over time, there may be consequences for the accumulation of MPs in terrestrial,
325
freshwater, or marine ecosystems derived from land-spreading of sewage sludge or bio-
326
solids.
327
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MPs entrained in biosolids which are applied to land, may be degraded through photo-
328
degradation and thermo-oxidative degradation.
52,53
exacerbating the problem of land-spread
329
MP pollution. The interaction of MPs with contaminants in the soil, could have major
330
consequences for the absorption and transportation of contamination elsewhere. Surface
331
weathering and the subsequent attachment of organic matter and the resulting negative charge
332
attracts metals including cadmium (Cd), lead (Pb) and zinc (Zn).
54
Whether agricultural land
333
is a sink or a source of MP pollution remains unclear. Microplastic fibres have been found on
334
land 15 years post application, and some evidence of vertical translocation through the soil
335
has also been found.
40
Possible impacts arising from land-applied MPs begin in the terrestrial
336
ecosystem with implications for terrestrial species such as earth worms
56
and birds feeding on
337
terrestrial ecosystems.
57
As legislation in the EU and the US generally permit the land
338
application of sewage sludge, there is a strong possibility that large amounts of MPs are
339
emitted to freshwater, where currently little is known on their impacts to species and
340
habitats.
58
Furthermore, buffer zones around freshwater bodies, which may be stipulated in
341
“codes of good practice”, do not take into account the mechanisms of transportation of MP
342
vertically through the soil or with surface runoff following a precipitation event. While
343
legislation currently takes into account pathogens as well as nutrient and metal concentrations
344
of treated sludge
59
it does not consider the presence of MPs within the sludge, and their
345
associated risks. The predicted exponential growth of the plastics industry for the coming
346
years
60
may be accompanied by a significant increase in MPs in the waste stream. Therefore,
347
vigilant management of cumulative sources of MPs such as sewage sludge or biosolids is
348
necessary. In particular, this study has highlighted the potential for treatment processes to
349
alter the counts of MPs which in-turn increases the available area for absorption/desorption of
350
organic pollutants
351
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Recommendations for further studies
353
Further investigations are required to investigate accelerated proliferation of MP pollution
354
through sludge treatment processes. In particular, the role of degradation by microorganisms
355
within the AD systems should be further investigated as a potential remediation method.
356
Knowledge gaps regarding the factors critical for the mobilisation and transport of MPs
357
which are likely to affect the pathway attenuation of land-spread sewage sludge MP pollution
358
should to be addressed in order to determine MP flow within the terrestrial system and to
359
freshwater systems. Only when the knowledge is acquired, can we estimate exposure and
360
associated risks to the environment from MP pollution.
361
362
Supporting Information
363
Detailed description of the dimensions of the elutriation column, accompanied by a
364
photograph and schematic representation. Flow rates and technique used for extraction of
365
MPs using the elutriation column are also included.
366
367
368
Acknowledgements
369
We acknowledge the technical assistance of Mark Deegan in construction of our Elutriation
370
system, Mark Croke and David James from Thermo Fisher Scientific UK for FTIR analyses
371
and the Environmental Protection Agency of Ireland for funding this research.
372
373
Bibliography
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report- Part I; Overview Report. European Commission Service Contract.
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No.070307/2008/517358/ETU/G4; 2013.
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report- Part I; Report on Options and Impacts. European Commission Service Contract
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sludge. Environ. Pollut. 2005, 138 (2), 201–211.
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Cole, M.; Webb, H.; Lindeque, P. K.; Fileman, E. S.; Halsband, C.; Galloway, T. S.
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(51) Richards, B. K.; Steenhuis, T. S.; Peverly, J. H.; McBride, M. B. Effect of sludge-
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radiation on materials. J. Photochem. Photobiol. B Biol. 1998, 46 (1-3), 96–103.Gu, J.-D.
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containing a pro-oxidant/pro-degradant additive. Polym. Degrad. Stab. 2013, 98 (11), 2117–
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2124.
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(54) Turner, A.; Holmes, L. A. Adsorption of trace metals by microplastic pellets in fresh
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water. Environmental Chemistry. 2015, pp 600–610.
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(55) Peyton, D.P., Healy, M.G., Fleming, G.T.A., Grant, J., Wall, D., Morrison, L.,
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Cormican, M., Fenton, O. Nutrient, metal and microbial loss in surface runoff following
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treated sludge and dairy cattle slurry application to an Irish grassland soil. Sci Total Environ.
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(56) Huerta Lwanga, E.; Gertsen, H.; Gooren, H.; Peters, P.; Salánki, T.; van der Ploeg, M.;
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Besseling, E.; Koelmans, A. A.; Geissen, V. Microplastics in the Terrestrial Ecosystem:
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Implications for Lumbricus terrestris (Oligochaeta, Lumbricidae). Environ. Sci. Technol.
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(57) Zhao, S.; Zhu, L.; Li, D. Microscopic anthropogenic litter in terrestrial birds from
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Shanghai, China: Not only plastics but also natural fibers. Sci. Total Environ. 2016, 550,
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1110–1115.
534
(58) Eerkes-Medrano, D.; Thompson, R. C.; Aldridge, D. C. Microplastics in freshwater
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of research needs. Water Res. 2015, 75, 63–82.
537
(59) Healy, M. G.; Fenton, O.; Forrestal, P. J.; Danaher, M.; Brennan, R. B.; Morrison, L.
538
Metal concentrations in lime stabilised, thermally dried and anaerobically digested sewage
539
sludges. Waste Manag. 2016, 48, 404–408.
540
(60) Plastics Europe. An Analysis of European plastic production, demand and waste data for
541
2011. Brussels, Belguim, 2012.
542
543
544
Table 1. Characteristics of municipal wastewater treatment sites investigated (adapted from
545
Healy et al., 2016)
546
Site
WWTP/
agglomeration
size (PEs)
leachate as %
of influent
BOD load
Industrial, and domestic
septic tank sludge
1
as %
of influent BOD load
Type of treatment
1A
2,362,329
<0.01
<0.01
Thermal drying, anaerobic
digestion
1B 284,696 0.3 24 Thermal drying
2 179,000 unknown 30 Anaerobic digestion
3
130,000
unknown
0.008
Thermal drying
4
101,000
2.0
unknown
Lime stabilisation
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5 31,788 0.25 unknown Lime stabilisation
6 25,000 0.7 0 Thermal drying
7
6,500
Unknown
Unknown
Thermal drying
1
Most recent available figures in all WWTPs (2013)
547
548
549
Table 2.
Particle size fraction (g) of lime stabilised (LS), anaerobically digested (AD) and
550
thermally dried (TD) samples (40 g).
551
Size fraction
Treatment type
LS AD TD
> 212 µm 3.004 ± 0.550 31.753± 0.578 35.503± 0.661
> 63 µm 27.410± 0.840 7.948± 0.7778 3.593± 0.894
> 45 µm 9.400± 1.166 0.327± 0.241 0.930± 0.486
< 45 µm 0.200 ± 0.213 0.000± 0.00 0.000± 0.000
552
553
554
555
556
557
558
Table 3. Breakdown of types of average microplastic abundance kg
-1
(dry matter) among
559
sites.
560
Microplastic Types
Site no. Treatment Fibres Fragments Films Spheres other
1A TD 9,113 511 255 89 44
1B AD 2,065 611 67 0 0
2 TD 5,583 588 222 44 67
3 AD 4,007 855 111 33 150
4 TD 13,675 1,143 366 33 178
5
LS
10,778
3,075
122
11
78
6 LS 4,762 5,228 11 0 11
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7
TD
3,463
511
167
0
56
Total - 53,447 12,521 1,321 211 583
% - 78.5 18.4 1.9 0.3 0.9
561
562
563
564
Figure 1. Average abundances and corresponding population equivalents of microplastics at 7
565
sites. Sites sharing the same letter are not significantly different (Mann- Whitney-U test, p >
566
0.005)
567
568
569
570
Figure 2. Stereomicrograph of mircoplastics fibres (A), other (B) and fragment (C).
571
572
573
574
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576
Figure 3. Fourier Transform Infrared Spectroscopy (FTIR) spectra within specimen
577
photographs of polyamide, polypropylene and Polyethylene terephthalate (PET).
578
579
580
581
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Figure 4. Abundance of microplastics in different size classes (A: 250-400 µm, B: 400-600
582
µm, C: 600-1000 µm, D: 1000-4000 µm) as a function of treatment type.
583
584
585
586
587
588
589
590
591
592
593
594
595
Figure 5. Diversity in morphology and surface texture of microplastics isolated from treated
596
sewage sludge. Scanning electron micrographs of fibrous particle from thermally dried (TD)
597
biosolids (A-C). Multi fibrous particle from lime stabilised (LS) biosolids (D-F). Overview of
598
non-fibrous particle from anaerobically digested (AD) biosolids (G-H). Presence of lamellae
599
or cleavage planes (arrow heads) on microplastic surface (I).
600
601
602
603
604
605
606
607
608
609
610
611
612
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Figure 6. Morphological and surface texture comparison between pre-treatment high density
613
polyethylene (HDPE) and HDPE particles isolated from treated sewage sludge. Scanning
614
electron images of pre-treatment HDPE (A-C) showing smooth non-degraded surface.
615
Scanning electron micrographs of HDPE particle from lime stabilised (LS) biosolids (D-F)
616
showing altered and weathered surface texture. Scanning electron micrograph of HDPE
617
particle from thermally dried (TD) biosolids (G-I) with evidence of blistering effect (arrow
618
heads) on polymer surface (I).
619
620
621
622
623
624
625
626
627
Figure 7. Morphological and surface texture comparison between pre-treatment polyethylene
628
(PE) and PE particle isolated from sewage sludge. Scanning electron images of pre-treatment
629
PE (A-C) with unaltered surface. Scanning electron micrographs of PE particle from
630
thermally dried (TD) biosolids (D-F) showing wrinkling and fracturing of polymer surface.
631
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... This includes pharmaceuticals, steroids, hormones, persistent organic pollutants (e.g. POP's), pathogenic microorganisms (Alvarenga et al., 2015) and microplastics (Mahon et al., 2016;. ...
... What's more, the type of microplastics in biosolids may be impacted by treatment type. Mahon et al. (2016), reported a higher abundance of smaller particles in lime stabilisation treated biosolids, compared to other treatment types as a result of a shearing and melting of microplastics during thermal processing, resulting in smaller particles. This may have larger implications, as when the plastics are melted, they can release potentially toxic substances (Tawfik and Huyghebaert, 1998;Whitt et al., 2016). ...
Thesis
Microplastics are an environmental issue of global concern. Although they have been found in a range of environments worldwide, their contamination in the terrestrial environment is poorly understood, particularly in relation to their source and fate. One source of microplastics in soils of particular concern is biosolids. Biosolids are the solid by-product of the wastewater treatment and are commonly spread on agricultural land as a fertilizer, indicating a potential route for microplastics into terrestrial soils. The aim of this thesis was therefore to broaden the understanding of microplastic contamination in agricultural soils in relation to biosolid application. The lack of suitable methods for microplastic detection and quantification is a major obstacle for determining their concentrations in soil environments. Therefore, an experiment was carried out to determine the best methods for microplastic extraction based on soil characteristics. The efficiency of organic matter removal methods was measured. Soils with a range of particle size distribution and organic matter content were spiked with a variety of microplastic types and density separation methods were tested. The optimal organic removal method was found to be hydrogen peroxide with organic removal rates up to 93%. The recovery efficiency of microplastics was variable across polymer types. Overall, canola oil was shown to be the best method for density separation, however, efficiency was dependent on the amount of organic matter in the soil. This outcome highlights the importance of including matrix-specific calibration in future studies considering a wide range of microplastic types, to avoid underestimation of microplastic contamination. To understand the sources and fate of microplastics in agricultural soils, these tailored methods were used to extract microplastics from samples collected from agricultural soils in the River Test catchment area in the UK. Soils were collected from fields which had historical biosolid application and these were compared to a similar set of fields which had never received biosolid application during summer and winter. The mean microplastic concentration was high in both the biosolid treated fields (874 MP/kg) and the untreated fields (664 MP/kg) and a wide variety of polymers were found across sites. There was a lack of significant difference between treated and untreated soils suggesting the influence of other sources and environmental processes. Additionally, soil samples were collected from five separate fields over the course of a year, before and after biosolid application. Microplastic contamination was ubiquitous across these fields up to a maximum concentration of 7950 MP/kg. Despite previous reports of high concentrations of microplastics in biosolids, their concentrations in soils did not significantly increase after application of biosolids. This suggests that biosolids may not be the key influencing factor in microplastic soil concentrations and transport out of soil systems is likely through horizontal (run off) and lateral (percolation) routes. Agricultural soils may thus be acting as a vector for microplastics to freshwater systems and the wider environment. Overall, the results of this thesis suggest that biosolids, whilst are likely a contributor, are not the sole source of microplastics in agricultural soils. The importance of additional sources and pathways are explored, and the complexities of the soil environment are considered, suggesting the highly dynamic nature of soil environment may determine the variability in microplastic concentrations. The research presented here significantly increased the understanding of microplastic sources and fate in agricultural soil systems while highlighting directions for future soil microplastic research.
... In addition to MPs (,5 mm), which are deliberately produced for use in the production of personal care products and large plastic products, large-size plastics also break down into MPs (,5 mm) when exposed to various factors such as mechanical abrasion and UV exposure (Song et al. 2017). Many studies are reporting that MPs are found in drinking water (Wong et al. 2021), freshwater (Yahaya et al. 2022), seawater (Núñez et al. 2021), landfill leachate (Sun et al. 2021), sludge of WWTPs (Mahon et al. 2017), atmosphere (Dris et al. 2015), soil (Zhao et al. 2021), sediments (Yahaya et al. 2022), food (Diaz-Basantes et al. 2020), and the body of aquatic organisms (Núñez et al. 2021). ...
Article
Full-text available
Since wastewater treatment plants (WWTPs) cannot completely remove microplastics (MPs) from wastewater, WWTPs are responsible for the release of millions of MPs into the environment even in 1 day. Therefore, knowing the sources, properties, removal efficiencies and removal mechanisms of MPs in WWTPs is of great importance for the management of MPs. In this paper, firstly the sources of MPs in WWTPs and the quantities and properties (polymer type, shape, size, and color) of MPs in influents, effluents, and sludges of WWTPs are presented. Following this, the MP removal efficiency of different treatment units (primary settling, flotation, biological treatment, secondary settling, filtration-based treatment technologies, and coagulation) in WWTPs is discussed. In the next section, details about MP removal mechanisms in critical treatment units (settling and flotation tanks, bioreactors, sand filters, membrane filters, and coagulation units) in WWTPs are given. In the last section, the mechanisms and factors that are effective in adsorbing organic–inorganic pollutants in wastewater to MPs are presented. Finally, the current situation and research gap in these areas are identified and suggestions are provided for topics that need further research in the future.
... A new system transports wastewater and rainwater separately; wastewater is sent to the sewage treatment plant, and rainwater is sent directly to rivers. The pathways of MP from sewage treatment plants to the environment are different by the size of MP; small and light MP are emitted with treated water runoff from the plant into the river, and heavier MP precipitate into the sewage sludge, which is used for an agricultural soil amendment and fertilizer in landfills 45,46 . In fact, MPs were found in agricultural soil to which sewage sludge has been applied 11 . ...
Article
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Marine microplastics are one of the global environmental issues. The present study examined whether rubber tips of artificial sports fields could be marine microplastics. We observed the migration of rubber tips from the artificial turf field to the surrounding ditch connected to sewer pipes and then examined the ingestion of rubber tips using the goldfish Carassius auratus. The rubber tips found in sediments in the ditch suggest that the rubber tips could be sent to the river and released into the ocean. The goldfish ingested rubber tips with or without fish feed, and rubber tips were found in the intestine. However, the fish discharged the rubber tips within 48 h after ingestion. These results indicate that ingestion of the rubber tips was not accidental but an active behavior. Therefore, artificial turf sports fields could be a source of marine microplastics and may cause hazardous effects on wild fishes through ingestion.
... The possibility of recycling these residual plastics is difficult because of practical challenges in addition to high economic costs (Steinmetz et al., 2016;Kasirajan and Ngouajio, 2012). Therefore, the soil has become the ultimate sink for residual plastics (Machado et al., 2018;Mahon et al., 2017;Liu et al., 2019). ...
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Environmental pollution of microplastics (MPs) is known to be anthropogenically mediated menace to biosphere and becoming a debatable concern globally. Large quantities of plastic fragments are left behind after crop cultivation. The leftover plastic debris, gradually degrade into minute fragments with a diameter of less than 5 mm, known as MPs. MPs are responsible for many changes in the soil physicochemical characteristics, including porosity, enzymatic activities, microbial activities, plant growth, and yield. Because of their ubiquitous nature, high specific surface area and strong hydrophobicity, MPs play an important role in the transportation of toxic chemicals such as plasticisers, polycyclic aromatic hydrocarbons (PAHs), antibiotics, and potentially toxic elements (PTEs). MPs may be transported deep into the soil and can pollute underground water. This review paper investigates the deleterious effects of MPs on the soil environment, enzymatic activities, soil microbes, flora, fauna and crop production, and highlights the general concept of MPs contamination as well as its possible environmental consequences. The review also converses some of the key areas for future research and for key stakeholders concerned with policymaking
... Thus, they are very likely to enter into the nearby rivers, and eventually into the marine (Blasing and Amelung 2018). Moreover, microplastics could also enter into farmlands through reclaimed water irrigation and sludge application (Mahon et al. 2017). Nevertheless, related environmental performance of primary microplastics are rarely reported. ...
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... Microplastics are plastic debris with <5 mm size that can enter the soil system through many pathways, prevalently the application of biosolids and sludges that intrinsically contain plastic particles (Mahon et al., 2017;Watteau et al., 2018). Microplastics may interact with other contaminants as metals and organic compounds through adsorption/desorption reactions contributing to pollutant transport and pollutant leaching (Davranche et al., 2019;Qi et al., 2020;Zhang et al., 2020;Chen et al., 2021). ...
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Plastic debris is widespread in the environment, but information on the effects of microplastics on terrestrial fauna is completely lacking. Here, we studied the survival and fitness of the earthworm Lumbricus terrestris (Oligochaeta, Lumbricidae) exposed to microplastics (Polyethylene, <150 μm) in litter at concentrations of 7, 28, 45, and 60% dry weight, percentages that, after bioturbation, translate to 0.2 to 1.2% in bulk soil. Mortality after 60 days was higher at 28, 45, and 60% of microplastics in the litter than at 7% w/w and in the control (0%). Growth rate was significantly reduced at 28, 45, and 60% w/w microplastics, compared to the 7% and control treatments. Due to the digestion of ingested organic matter, microplastic was concentrated in cast, especially at the lowest dose (i.e., 7% in litter) because that dose had the highest proportion of digestible organic matter. Whereas 50 percent of the microplastics had a size of <50 μm in the original litter, 90 percent of the microplastics in the casts was <50 μm in all treatments, which suggests size-selective egestion by the earthworms. These concentration-transport and size-selection mechanisms may have important implications for fate and risk of microplastic in terrestrial ecosystems.
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Cognisant of the negative debate and public sentiment about the land application of treated sewage sludges ('biosolids'), it is important to characterise such wastes beyond current regulated parameters. Concerns may be warranted, as many priority metal pollutants may be present in biosolids. This study represents the first time that extensive use was made of a handheld X-ray fluorescence (XRF) analyser to characterise metals in sludges, having undergone treatment by thermal drying, lime stabilisation, or anaerobic digestion, in 16 wastewater treatment plants (WWTPs) in Ireland. The concentrations of metals , expressed as mg kg À1 dry solids (DS), which are currently regulated in the European Union, ranged from 11 (cadmium, anaerobically digested (AD) biosolids) to 1273 mg kg À1 (zinc, AD biosolids), and with the exception of lead in one WWTP (which had a concentration of 3696 mg kg À1), all metals were within EU regulatory limits. Two potentially hazardous metals, antimony (Sb) and tin (Sn), for which no legislation currently exists, were much higher than their baseline concentrations in soils (17–20 mg Sb kg À1 and 23–55 mg Sn kg À1), meaning that potentially large amounts of these elements may be applied to the soil without regulation. This study recommends that the regulations governing the values for metal concentrations in sludges for reuse in agriculture are extended to include Sb and Sn.
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The permanent presence of microplastics in the marine environment is considered a global threat to several marine animals. Heavy metals and microplastics are typically included in two different classes of pollutants but the interaction between these two stressors is poorly understood. During 14 days of experimental manipulation, we examined the adsorption of two heavy metals, copper (Cu) and zinc (Zn), leached from an antifouling paint to virgin polystyrene (PS) beads and aged polyvinyl chloride (PVC) fragments in seawater. We demonstrated that heavy metals were released from the antifouling paint to the water and both microplastic types adsorbed the two heavy metals. This adsorption kinetics was described using partition coefficients and mathematical models. Partition coefficients between pellets and water ranged between 650 and 850 for Cu on PS and PVC, respectively. The adsorption of Cu was significantly greater in PVC fragments than in PS, probably due to higher surface area and polarity of PVC. Concentrations of Cu and Zn increased significantly on PVC and PS over the course of the experiment with the exception of Zn on PS. As a result, we show a significant interaction between these types of microplastics and heavy metals, which can have implications for marine life and the environment. These results strongly support recent findings where plastics can play a key role as vectors for heavy metal ions in the marine system. Finally, our findings highlight the importance of monitoring marine litter and heavy metals, mainly associated with antifouling paints, particularly in the framework of the Marine Strategy Framework Directive (MSFD).
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Treated municipal sewage sludge (“biosolids”) and dairy cattle slurry (DCS) may be applied to agricultural land as an organic fertiliser. This study investigates losses of nutrients in runoff water (nitrogen (N) and phosphorus (P)), metals (copper (Cu), nickel (Ni), lead (Pb), zinc (Zn), cadmium (Cd), chromium (Cr)), and microbial indicators of pollution (total and faecal coliforms) arising from the land application of four types of treated biosolids and DCS to field micro-plots at three time intervals (24, 48, 360 h) after application. Losses from biosolids-amended plots or DCS-amended plots followed a general trend of highest losses occurring during the first rainfall event and reduced losses in the subsequent events. However, with the exception of total and faecal coliforms and some metals (Ni, Cu), the greatest losses were from the DCS-amended plots. For example, average losses over the three rainfall events for dissolved reactive phosphorus and ammonium-nitrogen from DCS-amended plots were 5 and 11.2 mg L− 1, respectively, which were in excess of the losses from the biosolids plots. When compared with slurry treatments, for the parameters monitored biosolids generally do not pose a greater risk in terms of losses along the runoff pathway. This finding has important policy implications, as it shows that concern related to the reuse of biosolids as a soil fertiliser, mainly related to contaminant losses upon land application, may be unfounded.
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Treated municipal sewage sludge (“biosolids”) and dairy cattle slurry (DCS) may be applied to agricultural land as an organic fertiliser. This study investigates losses of nutrients in runoff water (nitrogen (N) and phosphorus (P)), metals (copper (Cu), nickel (Ni), lead (Pb), zinc (Zn), cadmium (Cd), chromium (Cr)), and microbial indicators of pollution (total and faecal coliforms) arising from the land application of four types of treated biosolids and DCS to field micro-plots at three time intervals (24, 48, 360 hr) after application. Losses from biosolids-amended plots or DCS-amended plots followed a general trend of highest losses occurring during the first rainfall event and reduced losses in the subsequent events. However, with the exception of total and faecal coliforms and some metals (Ni, Cu), the greatest losses were from the DCS-amended plots. For example, average losses over the three rainfall events for dissolved reactive phosphorus and ammonium-nitrogen from DCS-amended plots were 5 and 11.2 mg L-1, respectively, which were in excess of the losses from the biosolids plots. When compared with slurry treatments, biosolids generally do not pose a greater risk in terms of losses along the runoff pathway. This finding has important policy implications, as it shows that concern related to the reuse of biosolids as a soil fertiliser, mainly related to contaminant losses upon land application, may be unfounded.