Performance of an anaerobic membrane bioreactor for pharmaceutical wastewater treatment

Performance of an anaerobic membrane bioreactor for pharmaceutical wastewater treatment

Accepted Manuscript Performance of an Anaerobic Membrane Bioreactor for Pharmaceutical Wastewater Treatment Jan Svojitka, Luká š Dvoř ák, Martin Stude...

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Accepted Manuscript Performance of an Anaerobic Membrane Bioreactor for Pharmaceutical Wastewater Treatment Jan Svojitka, Luká š Dvoř ák, Martin Studer, Jürg Oliver Straub, Heinz Frömelt, Thomas Wintgens PII: DOI: Reference:

S0960-8524(17)30042-1 http://dx.doi.org/10.1016/j.biortech.2017.01.022 BITE 17517

To appear in:

Bioresource Technology

Received Date: Revised Date: Accepted Date:

14 October 2016 9 January 2017 11 January 2017

Please cite this article as: Svojitka, J., Dvoř ák, L., Studer, M., Oliver Straub, J., Frömelt, H., Wintgens, T., Performance of an Anaerobic Membrane Bioreactor for Pharmaceutical Wastewater Treatment, Bioresource Technology (2017), doi: http://dx.doi.org/10.1016/j.biortech.2017.01.022

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1

PERFORMANCE OF AN ANAEROBIC MEMBRANE BIOREACTOR FOR

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PHARMACEUTICAL WASTEWATER TREATMENT

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Jan Svojitka1, Lukáš Dvořák2, Martin Studer3, Jürg Oliver Straub 3, Heinz Frömelt4,

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Thomas Wintgens1*

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and Arts Northwestern Switzerland, Gründenstrasse 40, CH-4132 Muttenz, Switzerland

Institute for Ecopreneurship, School of Life Sciences, University of Applied Sciences

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University of Liberec, Studentská 2, 461 17, Liberec 1, Czech Republic

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F.Hoffmann-La Roche Ltd, Grenzacherstrasse 124, CH-4070 Basel, Switzerland

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ProRheno AG, Grenzstrasse 15, CH-4057 Basel, Switzerland

Centre for Nanomaterials, Advanced Technologies and Innovation, Technical

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E-mail: [email protected], [email protected], [email protected],

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[email protected], [email protected], [email protected]

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* Corresponding author

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1

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Abstract

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Anaerobic treatment of wastewater and waste organic solvents originating from the

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pharmaceutical and chemical industries was tested in a pilot anaerobic membrane

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bioreactor, which was operated for 580 days under different operational conditions. The

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goal was to test the long-term treatment efficiency and identify inhibitory factors. The

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highest COD removal of up to 97% was observed when the influent concentration was

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increased by the addition of methanol (up to 25 g·L–1 as COD). Varying and generally

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lower COD removal efficiency (around 78%) was observed when the anaerobic

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membrane bioreactor was operated with incoming pharmaceutical wastewater as sole

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carbon source. The addition of waste organic solvents (>2.5 g·L–1 as COD) to the

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influent led to low COD removal efficiency or even to the breakdown of anaerobic

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digestion. Changes in the anaerobic population (e.g., proliferation of the genus

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Methanosarcina) resulting from the composition of influent were observed.

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Keywords:

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Anaerobic membrane bioreactor (AnMBR); COD removal; pharmaceutical wastewater;

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fouling; biogas; inhibition of methanogens

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1 Introduction

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Industrial wastewater treatment by anaerobic biological processes is a proven

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technology with several advantages compared to aerobic treatment, such as production

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of biogas, lower energy costs and low excess sludge production (Lew et al., 2009).

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Although modern high-rate anaerobic reactors such as Up-flow Anaerobic Sludge

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Blanket (UASB) can achieve chemical oxygen demand (COD) removal efficiencies

2

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over 90% (Choi et al., 2013; Delforno et al., 2014), a more widespread use of anaerobic

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wastewater treatment is hampered by higher residual effluent pollution (Chen et al.,

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2008) and poor retention of biomass in the reactor (Lin et al., 2013). Furthermore, the

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granular biomass in high-rate anaerobic reactors can be negatively affected by the

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characteristics of various industrial wastewaters.

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Anaerobic membrane bioreactors (AnMBR) present an attractive approach for the

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treatment of wastewater with a high content of COD and suspended solids, high salinity

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and large variations of flow and composition, even in the presence of fat, oil and grease

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or inhibitory compounds (Dereli et al., 2012; Diez et al., 2012; Lin et al., 2013).

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Moreover, the membrane in an AnMBR represents a barrier for slow-growing

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microorganisms with specialised degradation pathways, resulting in an increase in their

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activity. For instance, Tao et al. (2012) used an AnMBR for the retention of slow-

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growing Anammox microorganisms; installing a membrane in the system increased

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their activity 19 times.

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Despite the aforementioned advantages, there are still several drawbacks associated

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with AnMBR, especially the lower filterability of the biomass leading to lower filtration

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fluxes compared to aerobic MBRs (Lin et al., 2013).

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AnMBRs have been successfully applied for the treatment of various industrial

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wastewaters at both pilot and full scale. Most applications targeted wastewater from

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food processing (e.g. Spagni et al., 2010; Wijekoon et al., 2011) since this is in general

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highly biodegradable, contains high concentrations of organic matter and often high

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amounts of suspended solids (Liao et al., 2006; Lin et al., 2013). AnMBRs have been

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further used for treatment of pulp and paper industry wastewater (Gao et al., 2010;

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Savant et al., 2006), textile industry wastewater (dos Santos et al., 2007) or polymer

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synthesis effluents (Araya et al., 1999). However, there have been few studies dealing

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with the treatment of pharmaceutical wastewater in AnMBRs to date, particularly in

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pilot plants treating real wastewater or full scale installations (Dvořák et al., 2015).

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Most studies focusing on the treatment of real pharmaceutical wastewaters have been

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conducted either in aerobic MBRs or in lab-scale AnMBRs. For instance, Ng et al.

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(2014) tested the treatment of pharmaceutical wastewater in a lab-scale AnMBR and

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achieved COD removal of 14.7 – 47.2%. The low organic removal efficiencies were

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caused by high salinity and complex nature of the organics in the wastewater. In another

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study Ng et al. (2015) evaluated microbial communities and AnMBR performance

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treating wastewater from a pharmaceutical factory in a lab scale AnMBR and

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demonstrated a positive influence of halophilic organisms on the treatment efficiency.

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The present study focused on the performance of an AnMBR pilot plant, which was fed

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by real wastewater originating from pharmaceutical and chemical production. The

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treatment feasibility and efficiency of the industrial wastewater under anaerobic

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conditions, biogas production and process stability under highly varying composition

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was assessed. Furthermore, concentrated waste organic solvents from the production

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were tested as a co-substrate for the anaerobic process with regard to their degradability

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and possible inhibition properties. The goal of the study was to evaluate whether

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anaerobic treatment can be a reliable and economical addition or alternative to current

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aerobic wastewater treatment.

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2 Material and methods

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2.1 Anaerobic membrane bioreactor

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The AnMBR consisted of a bioreactor (50 L) and external cross-flow membrane unit

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(Figure 1). Before entering the bioreactor, the incoming raw wastewater was pumped

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into a buffer tank (20 L) in order to equilibrate concentration and flow peaks as well as

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to adjust the pH to 7.0–7.5. In some phases of operation (specified in chapter 2.3) an

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additional substrate, i.e. methanol or waste organic solvents, was added to the incoming

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wastewater. The daily influent into the bioreactor was between 10 and 30 L.

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The AnMBR was operated under mesophilic conditions (35–37 °C). Continuous

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recirculation between the bioreactor and filtration step facilitated mixing of the reactor

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and pressurization of two tubular membranes (TAMI Industries; Nyons, France). The

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membranes were 1-inch ceramic (ZrO2–TiO2) tubes with 8 channels (MWCO 50 kDa)

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and had a total area of 0.25 m2. Filtration was operated at a flux of 8.4 L·m–2·h–1. Since

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the permeate flow exceeded the desired inflow rate, part of the permeate was

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periodically recycled back to the bioreactor in order to reach the targeted hydraulic

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retention time (HRT). The excess sludge was removed discontinuously three times per

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week. The volume of excess sludge was adjusted to the incoming organic load, so that

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the concentration of total solids in the bioreactor was kept above 10 g·L–1. The solids

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retention time (SRT) was calculated as the ratio between the bioreactor volume and the

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sum of sludge volume extracted weekly; the resulting value was between 120 and 450 d.

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2.2 Industrial wastewater and waste organic solvents

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The AnMBR was placed indoors at the industrial wastewater treatment plant (WWTP)

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ProRheno AG in Basel (Switzerland), so that part of the incoming wastewater could be

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pumped into the bioreactor. The industrial wastewater originated from

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pharmaceutical/chemical production with the main constituents being methanol and

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ethanol (forming on average 80–90% of the organic load); the rest was a mixture of

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other organic solvents, e.g., tetrahydrofuran, dichloromethane, acetone, ethyl acetate,

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tripropylamine, acetonitrile, toluene, isopropanol, acetone or dimethylacetamide. The

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wastewater composition fluctuated greatly due to varying production with a COD

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between 0.55 and 10.6 g·L–1 (the average value was 4.20 g·L–1). The wastewater was

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mostly free of suspended solids.

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Beside the wastewater, significant amounts of waste organic solvents (WOS) with

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minor water content (max. 50%) arise from pharmaceutical production. WOS samples

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were tested for their anaerobic biodegradability; two batches (WOS 9 and WOS 18,

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compositions shown in Table 1) were selected and used as additional substrate for the

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AnMBR. The composition of WOS varied largely from one batch to another. In most

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cases methanol, ethanol and acetone were the main constituents; other components

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included acetonitrile, dichloromethane, ethyl acetate, tetrahydrofuran, isopropyl acetate,

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n-heptane and toluene.

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2.3 Phases of AnMBR operation

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The AnMBR was operated for a period of 580 days. The operation was divided into

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three phases differing in influent composition and organic loading rate (Table 2).

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Phase 1: The AnMBR was inoculated with a mixture of granular anaerobic sludge

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taken from a UASB reactor treating food industry wastewater (high starch content) and

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anaerobic sludge from a municipal WWTP digester. The two inocula were mixed in 1:1

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ratio based on total solids (TS) and diluted to give a concentration of 10 g·L–1. The

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AnMBR was then fed by the incoming industrial wastewater to test its treatability under

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realistic conditions and varying feed quality. The operation of the AnMBR was started

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by the adaptation of the biomass consisting of slow filling the reactor with wastewater

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in the first five days with minimum effluent. Afterwards the influent rate was increased

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from 10 to 30 L·d–1 (in 5 L·d–1 steps, each approx. 30 d) in order to increase the organic

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loading.

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Phase 2: To avoid variations in the wastewater composition, batches of wastewater

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(1 m3) were collected and used as influent, each for a period of several weeks. In total

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three batches of wastewater were used during Phase 2 with COD between 3.24 and 7.54

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g·L–1, methanol and ethanol being the main constituents. Methanol and WOS 9 (1–

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7 g·L–1 as COD) were added to the wastewater in order to increase organic loading in a

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controlled way and test the treatability of the WOS 9 as a co-substrate. In later stages of

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Phase 2 gradually increasing amounts of methanol were added to the wastewater in

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order to increase the incoming COD concentration and test the operation of the AnMBR

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at high organic loading. The influent rate of wastewater was kept constant at 14.3 L·d–1.

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Phase 3: In Phase 3 the feeding of the AnMBR was switched from the batches of

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wastewater back to continuous dosing of the incoming industrial wastewater. The COD

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concentration in the influent was increased by a constant amount of methanol and WOS

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18 in order to keep the organic loading high. The goal of this phase was to test the

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operation at generally high loading combined with variations in wastewater quality and

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test the treatability of WOS 18 as a co-substrate.

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2.4 Analytical methods

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The samples of influent and permeate were automatically collected and kept at 4°C until

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analysis. The concentrations of chemical oxygen demand (COD), total nitrogen,

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phosphate phosphorus and sulfides in influent and permeate samples were regularly

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measured using cuvette tests (Hach-Lange, Düsseldorf, Germany). Total solids (TS)

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were analyzed according to the Standard methods (APHA, 2012). Gas chromatography

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(5890 Series II, Hewlett-Packard, Palo Alto, CA, USA) was used to determine the

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methane content in the biogas samples and to analyze the concentration of organic acids

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in the liquid samples.

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2.5 FISH analyses

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The samples for the fluorescence in situ hybridization (FISH) analyses were fixed 1:1

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(vol./vol) with ethanol, absolute. The content of methanogenic Archaea was monitored

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using a FISH (Snaidr et al., 1997) test kit with standardized methodology and prepared

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solutions (VIT® Methanogenic bacteria, vermicon AG, Germany). The test kit detects

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methanogens affiliated with the Euryarchaeota, which cover the range of Archaea

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reported to be responsible for methanogenesis (Liu and Whitman, 2008). In situ

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hybridization was performed according to the manual of the VIT® Methanogenic

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bacteria test kit. In addition, a set of specific oligonucleotide probes was used to analyse

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the samples with respect to their content of main microbial phyla either participating in

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anaerobic digestion or being present in anaerobic environmental samples (Weinberger

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and Höötmann, 2011). For the identification of the Euryarchaeota detected, more

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specific oligonucleotide probes were used which target methanogens on class, family

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and even genus level (Raskin et al., 1994). Total cell counts were determined by

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filtration of samples on polycarbonate filter membranes, followed by using DAPI (4',6-

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diamidino-2-phenylindole) staining. Total viable cell counts were performed by

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hybridizing the same filters using a mixture of oligonucleotide probes specific for the

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domains Bacteria and Archaea (Amann et al., 1995, Daims et al., 1999).

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Hybridized cells were excited and examined under a Zeiss Axiostar plus fluorescence

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microscope. The percentage of area covered with target organisms in comparison to all

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viable cells was analysed in at least ten randomly selected captured fields and an

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average percentage of the populations was calculated. An eyepiece with crossline

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graticules was used for cell counts and calculation of absolute numbers.

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2.6 Respirometric tests

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The biodegradability of WOS and their single components was evaluated by

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respirometric degradation tests using the biomass samples from the AnMBR. The

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degradation tests of WOS 9 were carried out with two biomass samples taken at the end

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of Phase 1 (Day 166) and beginning of Phase 2 (Day 180) before the addition of WOS 9

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into the AnMBR. Later during Phase 2 another mixture of organic solvents (WOS 18)

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was collected and subjected to a new series of respirometric tests with five biomass

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samples taken between days 250 and 287. Degradation tests of the industrial wastewater

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collected at the same day as each biomass sample were carried out parallel to the tests

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with WOS mixtures.

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In order to assess possible inhibitory effects of the individual components of the

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wastewater and WOS, two additional series of respirometric tests were carried out using

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biomass samples from day 306 and 325 respectively. Ethanol, ethyl acetate, acetonitrile,

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tetrahydrofuran, dichloromethane, n-heptane and toluene were tested as the sole

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substrate (1.0 g·L–1 COD) and in mixtures (0.5 + 0.5 g·L–1 COD) with methanol, which

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served as reference substrate.

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The original methodology based on OECD guideline No. 311 (OECD, 2006) was

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modified as follows. A mixture (50 mL) of dilution medium, substrate, tested sludge

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and water was prepared to obtain a final COD concentration of 1.0 g·L–1 and TS of

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1.5 g·L–1. The dilution medium containing phosphate buffer and trace elements was

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prepared in accordance with OECD guideline No. 311. The flasks were incubated at

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35°C and pressure readings were collected every 30 minutes. At the end of the test, the

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methane content in the head-space was measured. Based on the data from pressure

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measurement and gas analyses, the amount of methane produced in each incubation

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flask was calculated using the ideal gas law. The specific methane production rate

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(SMPR) was calculated from the pressure evolution curve assuming a linear dependence

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between the amount of methane produced and pressure. The maximum SMPR was

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identified from each test and used as a measure for biomass activity with given substrate.

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3 Results and discussion

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3.1 Treatment efficiency of the AnMBR

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Phase 1

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The AnMBR was continuously fed by incoming industrial wastewater with flow rate

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increasing in steps from 10 to 30 L·d–1. The influent in this phase was characterized by

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highly varying COD concentrations (1.50–10.6 g·L–1; average 4.30 g·L–1). The effluent

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COD concentrations were between 0.240 and 1.70 g·L–1 (average 0.83 g·L–1) resulting

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in a COD removal efficiency between 44 and 94% (78% on average) (Figure 2a). Most

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of the effluent COD constituted non-degraded compounds from the original feed, since

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the concentration of organic acids did not exceed 90 mg·L–1 (max. 16% of total COD).

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The highest removal efficiencies above 90% were observed in the periods of volumetric

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loading rate above 2.5 g COD L–1·d–1 – with peak concentrations in the influent (day

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40–47 and day 66–73, COD concentrations 8.0–10.6 g·L–1) and at the end of phase 1,

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when the organic loading was elevated by a higher wastewater flow. Similarly, the

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lowest observed removal efficiency (44% on day 136) corresponded to the lowest COD

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feed concentration. In general, the achieved loading rate of the AnMBR was relatively

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low compared to values typically used in AnMBRs (2–15 g COD L–1·d–1) or

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conventional high rate anaerobic reactors (4–40 g COD L–1·d–1) (Liao et al., 2006; Lin

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et al., 2013), therefore the energy recovery of the full scale application would be low. In

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order to improve the economic efficiency of the process, the following phases focused

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on increasing the wastewater concentration by the addition of concentrated waste

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organic solvents from production (WOS).

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Phase 2

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After switching to batch-wastewater the organic loading rate of the AnMBR was

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increased compared to phase 1 by adding methanol to the influent (COD of the resulting

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mixture 5.4–7.6 g·L–1). These modifications did not disturb the operation and the

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AnMBR exhibited stable COD removal efficiency of 89–93% (Figure 2b). After 10

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days, methanol was replaced by the WOS 9 as additional substrate (COD of the

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resulting mixture 7.1–11.3 g·L–1). Following the addition of the WOS 9 the COD

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removal dropped to 19% within 8 days and other signs of process imbalance were

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observed – accumulation of organic acids (up to 480 mg·L–1 acetate) and subsequently a

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decreasing pH in the bioreactor. Consequently, the WOS 9 was replaced by methanol

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and reactor loading was decreased by reducing the amount of methanol added to the

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influent (resulting COD 3.6–4.2 g·L–1). These operational parameters were maintained

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for one month (until day 240), which led to successful recovery of the system.

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Since the operational parameters and composition of the wastewater was constant

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during phase 2, the inhibition can be attributed to the addition of WOS 9. The

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degradation tests of WOS 9 were carried out before the addition of WOS 9 into the

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AnMBR and they indicated lower SMPR compared to the industrial wastewater, but it

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did not lead to complete inhibition of methane generation (Table 3). This finding

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contradicted the operational experience of the AnMBR, where imbalance in methane

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production and changes in microbial population were observed following the dosing of

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WOS 9 (see also chapter 3.3). The reason might be the difference between

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concentration of the WOS 9 in the respirometric test (1 g·L–1 COD) and in the AnMBR

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(5–7 g·L–1 COD). Therefore, inhibition may only become visible at high concentrations.

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Between days 240 and 350 the COD of the influent mixture increased from 7 to 44 g·L–

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1

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loading of the bioreactor rose from 1.8 to 12.7 g COD L–1·d –1. COD removal efficiency

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after the recovery was mostly above 90%, except for two short disturbances around days

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260 and 350. These probably resulted from too rapid an increase in incoming COD load

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but the reactor performance rapidly stabilized (Figure 2b).

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The results obtained during phase 2 confirmed that anaerobic treatment process of the

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tested wastewater can be operated efficiently and at high energy yield, provided a well

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degradable, non-inhibited co-substrate is available in a sufficient amount.

by the addition of methanol (up to 37 g·L–1 as COD); correspondingly the organic

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Phase 3

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Reactivated pumping of wastewater from the sewer directly to the AnMBR caused

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variations in wastewater concentration but they were dampened by constant addition of

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methanol (25 g·L–1 as COD) in the first 40 days of phase 3 and did not affect the COD

12

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removal efficiency (91 – 97%, Figure 2c). Starting from day 409 WOS 18 was added to

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the wastewater (8 g COD L–1 of WOS 18 + 17 g COD L–1 of methanol), which resulted

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in rapid deterioration of effluent quality and the dosing of WOS 18 was interrupted.

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After recovery of the degradation process the dose of WOS 18 was increased in steps

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(1.25 – 5.0 g COD L–1). A rapid drop of COD removal to 60% was observed when the

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dose of WOS 18 reached 5.0 g COD L–1. The operation continued with the addition of

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2.5 g COD L–1 WOS 18, leading to varying treatment efficiency (67 – 93%). A third

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drop in COD removal was observed between days 500 and 511 which was not preceded

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by an increase in WOS 18 addition. However, the organic loading rate between days

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480 and 495 underwent vigorous changes resulting from variations in wastewater

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concentration connected with changes in the dosing rate of external substrates, which

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might have induced additional stress on the biomass and resulted in a massive

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accumulation of organic acids up to 1000 mg·L–1 on Day 502. Additionally, a failure of

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the temperature control system occurred on day 515, when the bioreactor was

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overheated to 45°C. After this event, generally low COD removal efficiency (< 75%)

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was observed although the addition of WOS was stopped. The FISH analysis (chapter

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3.3) of the biomass sample from day 579 confirmed the collapse of the methanogenic

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population and the operation of the pilot unit was terminated.

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Although the respirometric tests showed substantially better biodegradability of WOS

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18 compared to WOS 9 (Table 3), low performance of the AnMBR was observed when

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the dose of this mixture reached 8.0, 5.0 and 2.5 g·L–1 COD. Again, this may be due to

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higher concentration used in the pilot experiment compared to the respirometric test.

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Unlike phase 2 the composition of wastewater in phase 3 varied, which might have had

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an additional influence on performance of the bioreactor besides the addition of WOS

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18. The disturbance in COD removal around day 415 (WOS dose 8.0 g·L–1 COD),

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coincided with an elevated concentration of tetrahydrofuran in the industrial wastewater

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(influent concentration 800 mg·L–1). This is however below the reported concentration

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of 11.5 g·L–1 leading to significant inhibition of anaerobic process (Yao et al., 2010).

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Elevated levels of N,N-dimethylacetamide (up to 2500 mg·L–1) were present in the

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influent during the second disturbance in COD removal around day 460, which is

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comparable to the value (>2000 mg·L–1) reported as having a negative effect on

315

microorganisms (EC10 value for activated sludge and EC50 for bacteria, respectively)

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(Merck KGaA, 2013). Even though the presence of more than 2000 mg·L–1 N,N-

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dimethylacetamide spanned a broader period than the observed disturbance, a

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combination with the influence of the WOS 18 cannot be excluded.

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Overall, the results showed that the WOS samples tested are not suitable as a co-

320

substrate for the full-scale treatment of industrial wastewater, since they can lead to

321

unpredictable breakdown of the process.

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3.2 Organic acids

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The total concentration of organic acids in the effluent from AnMBR was usually below

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150 mg·L–1, with acetate being the dominant acid. Accumulation of organic acids

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indicating anaerobic digestion imbalance was observed on a few occasions as a result of

327

the addition of WOS 9 (days 190–210), addition of WOS 18 combined with loading

328

variations (days 498–510) and reactor overheating (days 515–530) (Figure 3). Low

329

concentration of organic acids < 70 mg·L–1 was observed during the increasing loading

330

between days 230 and 400, even during the peak volumetric loading of 12 g COD L–

331

1

·d–1 and two disturbances in COD removal around days 260 and 350. This may be due

14

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to a low share of acetoclastic metabolism in the bioreactor: Bhatti et al. (1996)

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identified three main pathways governing the fate of methanol during anaerobic

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digestion – acetoclastic methanogenesis through formation of acetate, hydrogenotrophic

335

methanogenesis through formation of hydrogen and direct methylotrophic conversion

336

into methane by Methanosarcina type species. The FISH analyses of biomass samples

337

confirmed that Methanosarcina were the dominant type of methanogenic population in

338

the AnMBR, although the type of metabolism could not be identified by this analysis,

339

since the group can use a variety of substrates and metabolic pathways.

340

Although the amount of methane formed by the methylotrophic metabolism does not

341

usually exceed 10% (Jiang et al., 2009), the low accumulation of organic acids during

342

the breakthrough events suggests that this type of metabolism might have prevailed in

343

the specialized biomass established after inhibition at the beginning of Phase 2, which

344

was fed with a uniform substrate.

345

On day 515 the bioreactor overheated to 45 °C due to failure of the temperature control

346

system. The elevated temperature led to another peak in concentration of organic acids

347

(up to 400 mg·L–1). The concentration of organic acids returned to below 200 mg·L–1

348

soon after the repair of the temperature control, however it remained at an elevated level

349

compared to previous phases (around 150 mg·L–1) until the end of the operation.

350 351

3.3 FISH analyses

352

Ten biomass samples from the AnMBR collected during phases 2 and 3 were analyzed

353

by FISH. The results showed a massive drop in methanogenic organisms between day

354

173 and day 201, which framed the period with WOS 9 dosing. This corresponded to

355

the breakdown observed in methane production and indicated a toxic effect of this

15

356

substrate towards the methanogens. After the dosing of WOS 9 was stopped (day 193),

357

the counts of methanogens recovered and in the sample from day 229 reached the

358

original level (Figure 4). A population shift within the methanogenic organisms was

359

observed. While the first sample contained a mixture (around 50:50) of rod-shaped and

360

cluster-forming methanogens, both types (especially the rod-shaped organisms) were

361

suppressed in samples from days 194 and 201 and starting from day 215 only the

362

cluster-forming methanogens were present in the biomass. Genus-specific probes

363

applied in samples from day 327 on identified this population as Methanosarcina. The

364

genus Methanosarcina is a common methanogen in anaerobic digesters. Several studies

365

found them to be a dominant methanogen, e.g., in reactors fed by sewage sludge (Liu et

366

al., 2016) or pharmaceutical wastewater (Chelliapan et al., 2011). Due to the versatility

367

of their metabolism and overall robustness they also dominate in anaerobic bioreactors

368

under transition conditions (De Vrieze et al., 2012).

369

The counts of methanogens roughly followed the organic loading trend of the reactor

370

and reached their peak on day 327. The last sample from day 579 contained a very low

371

share of viable cells (6% compared to 27–42% in the remaining samples) including

372

methanogens. This corresponds to the poor reactor methane production observed in that

373

period.

374

Changes within the population of other microbial groups in the biomass samples were

375

also observed during the period with the dosing of WOS 9. Whereas Firmicutes,

376

Crenarchaeota and Actinobacteria disappeared or their counts were suppressed during

377

the addition of WOS, Chloroflexi, Cytophaga-Flexibacter subphylum and

378

Alphaproteobacteria were found in higher amounts at the end of the first sampling

379

campaign. This indicates that the addition of WOS 9 affected only some microbial

16

380

groups, whereas the microorganisms less sensitive to the substances present rapidly

381

occupied the vacant habitat. The disappearance of Firmicutes and Actinobacteria, which

382

are connected with acidogenesis and acetogenesis (Levén et al., 2007), was remarkable.

383

This was probably the result of the lack of suitable substrates, since the AnMBR

384

influent contained few organic compounds other than methanol at that time.

385

In samples from day 327 onwards, Firmicutes and other bacteria were detected again in

386

gradually increasing numbers, presumably because more diversified substrate was

387

added to the AnMBR in phase 3. Relatively few viable cells were found in the last

388

sample from day 579 (after reactor overheating) with Firmicutes forming 22% of all

389

cells. Such an evolution is usually a sign of a collapsing reactor performance needing

390

new inoculation, which was also the case in this study.

391

A decreasing trend was observed for hydrolysis bacteria (Chloroflexi and Cytophaga-

392

Flexibacter-subphylum) after the population shift (from day 215 on), presumably due to

393

the specific type of substrates in the reactor, which had low hydrolytic requirements.

394 395

3.4 Biogas production and quality

396

The biogas produced contained about 40–70% methane, on average about 60%. The

397

amount of biogas formed was between 0.5 and 0.6 L·g–1 COD removed throughout all

398

phases of operation, corresponding to 0.30–0.36 L methane per g COD. Based on the

399

equilibrium calculation of known sulfide content in the effluent, the hydrogen sulfide

400

concentration in the gas was mostly below 0.03 g·m–3, but would reach approximately

401

14 g·m–3 (1% v/v) during the peak sulfide concentration detected.

402 403

3.5 Total solids and membrane performance

17

404

The total solids (TS) in the bioreactor roughly followed the trend of the organic loading

405

rate and varied between 9 and 23 g·L–1 (Figure 5). Despite faster sludge wasting

406

(corresponding SRT approx. 120 d) the concentration of TS roughly doubled from 11 to

407

23 g·L–1 during phase 2. The effect of accentuated biomass growth resulting from

408

higher organic loading probably combined with the introduction of suspended solids

409

(approx. 0.5 g·L–1) contained in the batch of wastewater used between days 290 and 366.

410

The membrane was presumably carrying a layer of foulants from the previous

411

experiments (data not shown) and a rapid increase in TMP was observed following the

412

inoculation so that chemical cleaning was necessary after 90 days of operation. In the

413

following more than 300 days only a minor decrease of permeability was observed

414

without any chemical cleaning of the membrane and despite the rapid increase of TS.

415

No fouling problems were observed up to 25 g·L–1 TS (Figure 5).

416

On the other hand, the membrane permeability decreased in the last 80 days of phase 3,

417

triggered by difficulties with maintaining the desired cross-flow velocity (3 m·s–1)

418

rather than by changes in the structure of the mixed liquor. Since the AnMBR had no

419

automatic control of the retentate flow, the filtration was repeatedly operated at low

420

cross-flow velocities close to dead-end mode for a period of several hours. This

421

probably led to accumulation of sludge in the membrane channels, as revealed by the

422

inspection of membranes at the end of the operation.

423 424

3.6 Degradation tests with the components of WOS

425

In the two series of degradation tests carried out with the components of the WOS, pure

426

methanol was the substrate with highest degradation rate (SMPR) (Figure 6), due to the

427

pre-adaptation of the biomass to this substrate. Degradation rate was substantially lower

18

428

when methanol was mixed with other substrates including well degradable organic

429

compounds (ethanol, ethyl acetate). Dichloromethane was the only substrate which

430

completely inhibited anaerobic degradation in mixture with methanol at the

431

concentration tested (0.5 g·L–1 COD, i.e. 0.88 g·L–1). This concentration was

432

comparable to that in the AnMBR influent during phase 2 (530 mg·L–1) and phase 3

433

(110 mg·L–1) and higher than the 50% inhibition concentration (IC50) in a UASB reactor

434

reported by Ozdemir et al. (2010) – 43 mg·L–1. Thus, dichloromethane could be

435

responsible for the observed breakdowns in COD removal in the AnMBR.

436

A relatively low degradation rate was observed for toluene. Toluene might have played

437

a more important role in the inhibition during phase 2 when the dosing of WOS 9 raised

438

its concentration in the influent to 160 mg·L–1. The same amount was present in the

439

respirometric test with low SMPR when degrading the toluene/methanol mixture and

440

320 mg·L–1 in the test with pure toluene. Inhibition (IC50) of anaerobic digestion was

441

reported in a similar range between 146 and 1000 mg·L–1 by Orellana (2000) and

442

Kayembe (2013). On the other hand, the concentration of toluene during the dosing of

443

WOS 18 did not exceed 30 mg·L–1.

444

In one test the degradation of acetonitrile mixed with methanol (Figure 6a) was low

445

compared to other substrates. However, acetonitrile presumably had no effect on the

446

breakdown observed in AnMBR performance since its concentration in the influent was

447

well below the dose in the degradation tests and did not exceed 30 mg·L–1.

448

Besides organic compounds, some inorganic substances, e.g. free ammonia or high

449

concentrations of salts in general, are known to inhibit anaerobic digestion (Chen et al.,

450

2008). However, they presumably had a low effect on the inhibition of COD removal in

451

the AnMBR, since the total nitrogen concentration in the bioreactor (< 300 mg·L–1) was

19

452

below the reported inhibitory limits of 1700 mg N-HH4 + L–1 (Yenigün and Demirel,

453

2013), while the salt content of 0.7 – 6 g·L–1 in the AnMBR was below the inhibitory

454

values of >13 g·L–1 reported for non-acclimated sludge (Lefebvre et al., 2007; Lefebvre

455

and Moletta, 2006).

456

Overall, it was not clear why the WOS samples, which showed reasonable

457

biodegradability in the single tests, had such a detrimental influence on the performance

458

of the AnMBR. Besides the influence of the solvents discussed above, other

459

components that are not detectable by GC – analysis in the WOS might also be

460

responsible for the negative effects. For a commercial operation, it would be important

461

to evaluate this in more detail, because the use of WOS is a key point in increasing the

462

biogas yield as well as in lowering the disposal cost for WOS.

463 464

4 Conclusions

465

The industrial wastewater alone showed good anaerobic degradability in the AnMBR,

466

however its low concentration and varying composition decreased the treatment

467

efficiency. Constant and high removal efficiency (>90%) was reached when methanol

468

was added to the influent and variations of its composition were limited. Acidification,

469

poor methane production and shifts in methanogenic population were experienced when

470

waste organic solvents were added to the influent as co-substrate. Therefore, a

471

commercially viable anaerobic treatment of the current wastewater and WOS mixtures

472

seems not possible at the moment. Well-controlled cross-flow conditions in the

473

membrane channels were necessary for stable filtration and low fouling.

474 475

Acknowledgements

20

476

We are very grateful to F. Hoffmann-La Roche for financial support, experiments and

477

access to analytical results. ProRheno AG is kindly thanked for kind cooperation and

478

support. Thanks also to students and technicians of the FHNW for pilot plant

479

supervision, analytics and technical support. Project LO1201 „National Programme for

480

Sustainability I“, and the OPR&DI project No. CZ.1.05/2.1.00/01.0005 is also

481

acknowledged.

482 483

References

484

1.Amann RI, Ludwig W, Schleifer KH. 1995. Phylogenetic identification and in situ

485

detection of individual microbial cells without cultivation. Microbiol Rev. 59(1):143-69.

486

2.APHA, 2012. Standard Methods for the Examination of Water and Wastewater, 22

487

edition. ed. American Water Works Assn, Washington, DC.

488

3.Araya, P., Aroca, G., Chamy, R., 1999. Anaerobic treatment of effluents from an

489

industrial polymers synthesis plant. Waste Management 19, 141–146.

490

doi:10.1016/S0956-053X(99)00013-6

491

4.Bhatti, Z.I., Furukawa, K., Fujita, M., 1996. Feasibility of methanolic waste treatment

492

in UASB reactors. Water Research 30, 2559–2568. doi:10.1016/S0043-1354(96)00144-

493

3

494

5.Chelliapan, S., Wilby, T., Yuzir, A., Sallis, P.J., 2011. Influence of organic loading on

495

the performance and microbial community structure of an anaerobic stage reactor

496

treating pharmaceutical wastewater. Desalination 271, 257–264.

497

doi:10.1016/j.desal.2010.12.045

21

498

6.Chen, Z., Ren, N., Wang, A., Zhang, Z.-P., Shi, Y., 2008. A novel application of

499

TPAD–MBR system to the pilot treatment of chemical synthesis-based pharmaceutical

500

wastewater. Water Research 42, 3385–3392. doi:10.1016/j.watres.2008.04.020

501

7.Choi, W.-H., Shin, C.-H., Son, S.-M., Ghorpade, P.A., Kim, J.-J., Park, J.-Y., 2013.

502

Anaerobic treatment of palm oil mill effluent using combined high-rate anaerobic

503

reactors. Bioresource Technology, Challenges in Environmental Science and

504

Engineering (CESE-2012) 141, 138–144. doi:10.1016/j.biortech.2013.02.055

505

8.Daims, H., A. Brühl, R. Amann, K.-H. Schleifer, and M. Wagner. 1999. The domain-

506

specific probe EUB338 is insufficient for the detection of all Bacteria: development and

507

evaluation of a more comprehensive probe set. Syst. Appl. Microbiol. 22:434–444.

508

9.De Vrieze, J., Hennebel, T., Boon, N., Verstraete, W., 2012. Methanosarcina: The

509

rediscovered methanogen for heavy duty biomethanation. Bioresource Technology 112,

510

1–9. doi:10.1016/j.biortech.2012.02.079

511

10.Delforno, T.P., Moura, A.G.L., Okada, D.Y., Varesche, M.B.A., 2014. Effect of

512

biomass adaptation to the degradation of anionic surfactants in laundry wastewater

513

using EGSB reactors. Bioresource Technology 154, 114–121.

514

doi:10.1016/j.biortech.2013.11.102

515

11.Dereli, R.K., Ersahin, M.E., Ozgun, H., Ozturk, I., Jeison, D., van der Zee, F., van

516

Lier, J.B., 2012. Potentials of anaerobic membrane bioreactors to overcome treatment

517

limitations induced by industrial wastewaters. Bioresource Technology, Membrane

518

Bioreactors (MBRs): State-of-Art and Future 122, 160–170.

519

doi:10.1016/j.biortech.2012.05.139

520

12.Diez, V., Ramos, C., Cabezas, J.L., 2012. Treating wastewater with high oil and

521

grease content using an Anaerobic Membrane Bioreactor (AnMBR). Filtration and

22

522

cleaning assays. Water Science and Technology 65, 1847–1853.

523

doi:10.2166/wst.2012.852

524

13.dos Santos, A.B., Cervantes, F.J., van Lier, J.B., 2007. Review paper on current

525

technologies for decolourisation of textile wastewaters: Perspectives for anaerobic

526

biotechnology. Bioresource Technology 98, 2369–2385.

527

doi:10.1016/j.biortech.2006.11.013

528

14.Dvořák, L., Gómez, M., Dolina, J., Černín, A., 2015. Anaerobic membrane

529

bioreactors—a mini review with emphasis on industrial wastewater treatment:

530

applications, limitations and perspectives. Desalination and Water Treatment 0, 1–15.

531

doi:10.1080/19443994.2015.1100879

532

15.Gao, W.J.J., Lin, H.J., Leung, K.T., Liao, B.Q., 2010. Influence of elevated pH

533

shocks on the performance of a submerged anaerobic membrane bioreactor. Process

534

Biochemistry 45, 1279–1287. doi:10.1016/j.procbio.2010.04.018

535

16.Jiang, N., Wang, Y., Dong, X., 2009. Methanol as the primary methanogenic and

536

acetogenic precursor in the cold zoige wetland at tibetan plateau. Microbial Ecology 60,

537

206–213. doi:10.1007/s00248-009-9602-0

538

17.Kayembe, K., 2013. Comparative study of benzyl and phenyl compounds chemical

539

structure effects on the inhibition of methane production by digested pig manure

540

methanogens. International Research Journal of Pure and Applied Chemistry 3, 48–58.

541

doi:10.9734/IRJPAC/2013/2489

542

18.Lefebvre, O., Moletta, R., 2006. Treatment of organic pollution in industrial saline

543

wastewater: A literature review. Water Research 40, 3671–3682.

544

doi:10.1016/j.watres.2006.08.027

23

545

19.Lefebvre, O., Quentin, S., Torrijos, M., Godon, J.J., Delgenès, J.P., Moletta, R., 2007.

546

Impact of increasing NaCl concentrations on the performance and community

547

composition of two anaerobic reactors. Applied Microbiology and Biotechnology. 75,

548

61–69. doi:10.1007/s00253-006-0799-2

549

20.Levén, L., Eriksson, A.R.B., Schnürer, A., 2007. Effect of process temperature on

550

bacterial and archaeal communities in two methanogenic bioreactors treating organic

551

household waste. FEMS Microbiology Ecology 59, 683–693. doi:10.1111/j.1574-

552

6941.2006.00263.x

553

21.Lew, B., Tarre, S., Beliavski, M., Dosoretz, C., Green, M., 2009. Anaerobic

554

membrane bioreactor (AnMBR) for domestic wastewater treatment. Desalination 243,

555

251–257. doi:10.1016/j.desal.2008.04.027

556

22.Liao, B.-Q., Kraemer, J.T., Bagley, D.M., 2006. Anaerobic membrane bioreactors:

557

applications and research directions. Critical Reviews in Environmental Science and

558

Technology 36, 489–530. doi:10.1080/10643380600678146

559

23.Lin, H., Peng, W., Zhang, M., Chen, J., Hong, H., Zhang, Y., 2013. A review on

560

anaerobic membrane bioreactors: Applications, membrane fouling and future

561

perspectives. Desalination 314, 169–188. doi:10.1016/j.desal.2013.01.019

562

24.Liu Y, Whitman WB. 2008. Metabolic, phylogenetic, and ecological diversity of the

563

methanogenic archaea. Ann N Y Acad Sci. 1125:171-89.

564

25.Liu, C., Li, H., Zhang, Y., Si, D., Chen, Q., 2016. Evolution of microbial community

565

along with increasing solid concentration during high-solids anaerobic digestion of

566

sewage sludge. Bioresource Technology 216, 87–94.

567

doi:10.1016/j.biortech.2016.05.048

568

26.Merck KGaA, 2013. N,N-Diethylacetamide for synthesis. Safety data sheet.

24

569

27.Ng K.K., Shi X., Tang M.K.Y., Ng H.Y. 2014. A novel application of anaerobic bio-

570

entrapped membrane reactor for the treatment of chemical synthesis-based

571

pharmaceutical wastewater. Separation and Purification Technology 132, 634–643.

572

doi:10.1016/j.seppur.2014.06.021

573

28.Ng, K.K., Shi, X., Ng, H.Y., 2015. Evaluation of system performance and microbial

574

communities of a bioaugmented anaerobic membrane bioreactor treating pharmaceutical

575

wastewater. Water Research 81, 311–324. doi: 10.1016/j.watres.2015.05.033

576

29.OECD, 2006. Test No. 311: Anaerobic Biodegradability of Organic Compounds in

577

Digested Sludge: by Measurement of Gas Production. OECD, Paris.

578

30.Orellana, R.I., 2000. Effects of various substrates on anaerobic biodegradation of

579

benzene and toluene. PhD thesis, The George Washington University, Washington, DC.

580

31.Ozdemir, C., Sen, N., Dursun, S., Kalipci, E., 2010. Removal of dichloromethane in

581

up-flow anaerobic sludge bed reactors and methane production. Asian Journal of

582

Chemistry 22, 6423–6436.

583

32.Raskin L, Stromley JM, Rittmann BE, Stahl DA. 1994. Group-specific 16S rRNA

584

hybridization probes to describe natural communities of methanogens. Appl Environ

585

Microbiol. 60(4):1232-40.

586

33.Savant, D.V., Abdul-Rahman, R., Ranade, D.R., 2006. Anaerobic degradation of

587

adsorbable organic halides (AOX) from pulp and paper industry wastewater.

588

Bioresource Technology 97, 1092–1104. doi:10.1016/j.biortech.2004.12.013

589

34.Snaidr J, Amann R, Huber I, Ludwig W, Schleifer KH. 1997. Phylogenetic analysis

590

and in situ identification of bacteria in activated sludge. Appl Environ Microbiol.

591

63(7):2884-96.

25

592

35.Spagni, A., Casu, S., Crispino, N.A., Farina, R., Mattioli, D., 2010. Filterability in a

593

submerged anaerobic membrane bioreactor. Desalination 250, 787–792.

594

doi:10.1016/j.desal.2008.11.042

595

36.Tao, Y., Gao, D.-W., Fu, Y., Wu, W.-M., Ren, N.-Q., 2012. Impact of reactor

596

configuration on anammox process start-up: MBR versus SBR. Bioresource

597

Technology 104, 73–80. doi:10.1016/j.biortech.2011.10.052

598

37.Weinberger G, Höötmann U. 2011. Wastewater treatment plants – a look into a black

599

box. IPW 11-12/2011:18-21. www.ipwonline.de.

600

38.Wijekoon, K.C., Visvanathan, C., Abeynayaka, A., 2011. Effect of organic loading

601

rate on VFA production, organic matter removal and microbial activity of a two-stage

602

thermophilic anaerobic membrane bioreactor. Bioresource Technology, Special Issue on

603

Challenges in Environmental Science and Engineering, CESE-2010: Technological

604

Advances in Waste Treatment for a Sustainable Future 102, 5353–5360.

605

doi:10.1016/j.biortech.2010.12.081

606

39.Yao, Y., Guan, J., Tang, P., Jiao, H., Lin, C., Wang, J., Lu, Z., Min, H., Gao, H.,

607

2010. Assessment of toxicity of tetrahydrofuran on the microbial community in

608

activated sludge. Bioresource Technology 101, 5213–5221.

609

doi:10.1016/j.biortech.2010.02.051

610

40.Yenigün, O., Demirel, B., 2013. Ammonia inhibition in anaerobic digestion: A

611

review. Process Biochemistry 48, 901–911. doi:10.1016/j.procbio.2013.04.012

612

26

613

Figure 1 Flow scheme of the pilot AnMBR.

614 615

27

616

Figure 2 COD concentration and COD removal during phase 1 (a), phase 2 (b) and

617

phase 3 (c). The areas in the diagram indicate the target dose of methanol (grey) and

618

WOS (yellow), both expressed as COD.

619

620

621 622 28

623

Figure 3 Organic acids in the effluent and organic loading of the AnMBR. The yellow

624

area in the diagram indicates target dose (expressed as COD) of WOS 9 (Phase 2) and

625

WOS 18 (Phase 3), respectively.

626 627

29

628

Figure 4 Comparison of biomass samples according to the absolute cell counts per mL

629

of each population.

630 631

30

632

Figure 5 Permeability (K) and total solids over AnMBR operation.

633 634

31

635

Figure 6 Maximum SMPR during respirometric tests of anaerobic biodegradability

636

with individual components of the WOS mixed with methanol (a) and as sole substrates

637

(b). Respirometric tests a) and b) were conducted with two different sludge samples.

638

MeOH – methanol, CH2Cl2 – dichloromethane, THF – tetrahydrofuran.

639

640 641

32

642

Table 1 Composition of WOS 9 and WOS 18 added to the pharmaceutical wastewater

643

during Phases 2 and 3. Parameter pH Water Methanol Ethanol Acetone Dichloromethane Ethyl acetate Tetrahydrofuran n-Heptane Toluene Ammonia nitrogen

WOS 9 (Phase 2) 11.3 14.7% 62.5% 6.4% 0.5% 6.9% 2.8% 0.4% – 2.1% 2.7%

WOS 18 (Phase 3) 10.9 6.0% 73.6% 2.0% 9.6% 1.0% 3.0% 4.4% 0.5% – –

644 645

33

646

Table 2 Technological parameters during different operational phases. Duration Wastewater Additional Phase [d] collection substrate 1

175

2

191

3

214

Sludge Volumetric HRT concentration loading [d] [g TS·L–1] [g COD·L–1·d–1]

no 7.3–13.1 co-substrate methanol, Batch 10.0–24.0 WOS 9 methanol, Continuous 18.3–26.4 WOS 18 Continuous

0.6–4.0

1.7–5

1.0–12.7

3.5

4.5–14.2

3–3.5

647 648

34

649

Table 3 Maximum SMPR during respirometric tests of anaerobic degradability with

650

WOS 9, WOS 18 and wastewater. Substrate

WOS 9

WOS 18

Wastewater

Maximum SMPR [mg COD g-1 h-1 ]

10.5 ± 3.5

37.9 ± 8.8

47.4 ± 6.0

651 652

35

653

Highlights

654

1.

Wastewater from pharmaceutical industry can be treated by anaerobic MBR technology.

655

2.

Varying wastewater composition had negative influence on the treatment efficiency.

656

3.

Addition of waste organic solvents caused inhibition of the anaerobic degradation.

657

4.

Dichloromethane exhibited the strongest inhibitory effect on anaerobic degradation.

658

5.

Changes in the biocenosis reflected the substrate pattern and inhibition events.

659 660

661

36