*Article* **Comparison of Di**ff**erent Biofilter Media During Biological Bed Maturation Using Common Carp as a Biogen Donor**

### **Mateusz Sikora \*, Joanna Nowosad and Dariusz Kucharczyk**

Department of Ichthyology and Aquaculture, University of Warmia and Mazury, Al. Warszawska 117A, PL 10-701 Olsztyn, Poland; nowosad.joanna@gmail.com (J.N.); darekk56@gmail.com (D.K.) **\*** Correspondence: sikora0404@gmail.com

Received: 3 December 2019; Accepted: 7 January 2020; Published: 15 January 2020

**Abstract:** This experiment analysed the operation of submerged and dripping biological filters with three types of filling: commercial fitting HXF12KLL (CF), two innovative polypropylene aggregates (PPA) and polyethylene screw caps for PET bottles (PSC). The experiment determined the time needed to reach full filter functionality at the maturation stage, the time needed to start successive stages of the nitrification process and the maximum concentration of each nitrogen compound in water in the recirculation systems. The filter operation characteristics after the maturation stage were also examined. These issues are crucial during the preparation and launch of new aquaculture facilities. A literature analysis indicated that the ability of biological filters to oxidise nitrogen compounds is affected by a number of factors. Studies conducted at various centres have covered selected aspects and factors affecting the effectiveness of biological filters. During this study, the model fish common carp (*Cyprinus carpio*) was used. The current experiment involved examination of biological filter maturation and operation during the carp fry rearing stage, which allowed the biofilter operation characteristics to be determined. At the third day of the experiment, the ammonium concentration reached approximately 3 mg NH4-N/dm3. It remained at this level for 10 days and later decreased below 0.25 mg NH4-N/dm3. The maximum nitrite concentration ranged from 11.7 mg/dm<sup>3</sup> to 20.9 mg NO2-N/dm<sup>3</sup> within 9 to 20 days and later decreased with time. Nitrate concentrations were seen to increase during the experiment. The all applied biofilter media showed possibility to be used in commercial aquaculture systems.

**Keywords:** biofilter maturation; nitrogen compounds; recirculating aquaculture system (RAS)

### **1. Introduction**

Annual fish and seafood consumption has been increasing steadily. The amount of fish and seafood obtained globally is limited and annual output has remained at 90 million tonnes for the past decade. For this reason, any increase in the amount of fish and seafood is associated with aquaculture and its dynamic growth. The production of aquaculture accounted for 44% of the total output in 2014 [1,2]. Installations used for animal production can be classified in regard to the degree of water recirculation: (1) open/flow-through systems, in which water is used once; (2) semi-open/semi-closed systems, in which water is used multiple times before being removed from the system, and (3) closed systems, in which only water loss is replenished. Higher degrees of water recirculation require more complex systems to purify it. RAS denotes technologies of repeated water reuse in a closed system. However, water loss in breeding systems needs to be replenished for multiple reasons, not only because of evaporation. To minimise the need for replenishing water loss, RAS systems are equipped with complex water treatment and purification systems, including mechanical and biological filters, UV sterilisation devices, water ozonation systems and others [3,4]. The most important of

these systems are biological filters, which control toxic nitrogen species and are necessary for RAS operation [3–5]. The fast increasing of aquaculture production in RAS is not possible without new developed technologies. One of the main focuses is media for biofilters, upon which the effectiveness of nitrification is dependant. This also an influence of biofilter costs, so new and cheaper biofilter media are necessary to involve in further aquaculture production growth [1–5].

The removal of toxic nitrogen from water is affected by nitrification, consisting of the biological oxidation of ammonium nitrogen to nitrite (III) nitrogen, followed by oxidation of the latter to nitrate (V) nitrogen. A crucial role in the nitrification process is played by Nitroso- (oxidation of ammonium to nitrite) and Nitro- (oxidation of nitrite to nitrate) nitrification bacteria [6,7]. Total ammonium nitrogen (TAN) is one of the major limiting factors in the design and operation of RAS systems [8] and is one of the metabolites which is formed in the digestion of proteins and the transformation of amino acids [9] given to fish with feed. It occurs as a sum of two species: dissociated (ammonium ion, NH4 +) and undissociated (ammonia, NH3) [8,10]. Ammonium nitrogen in its dissociated form is relatively non-toxic, whereas the undissociated form is highly toxic [8,9,11–17]. For this reason, it must be controlled [17] and removed from the system or oxidised to a less toxic nitrogen form [8,18]. Ammonia levels in RAS systems are controlled by nitrification [17]. The NH4 +/NH3 ratio depends mainly on the environment's pH. The amount of toxic NH3 increases along with higher pH, depending on the temperature, pressure and salinity. Excessive amounts of ammonia lead to tremors in fish, coma and death [10]. The ammonia removal process is affected by a number of factors. The most important of them is the availability of oxygen (4.57 g of oxygen is needed to oxidise 1 g of ammonium nitrogen) and the rate of its diffusion into the biomembrane [8]. A decrease in water saturation with oxygen results in inhibition of ammonium nitrogen oxidation [19,20]. Important factors also include temperature, pH, salinity and organic matter burden on the biological filter [8]. Nitrite nitrogen are intermediate compounds in nitrification [6,7]. Due to their toxic effect on fish, they are an equally important factor in fish breeding as ammonium nitrogen [7,14,21,22] and are responsible for chronic diseases leading to fish death [7]. Nitrate nitrogen is the last step in the nitrification process [6,7] and has long been regarded as harmless. Recent studies have shown that it should be taken into account when optimising fish breeding. This nitrogen form is relatively harmless, but long-term exposure is suspected of having deadly toxic effects [21,23], although further studies are needed [23]. Analysis of the nitrification process gives the information if the biofilter is matured. In RAS equipped with such a biofilter, the level of ammonium and nitrite concentrations are low and stable [18]. It is especially important for finfish species, which are sensitive to toxic nitrogen compounds, like salmonids [4,6,7,11,12].

The kinetics of the reaction were not studied during the experiment, because the purpose of the study was to examine whether the applied experimental fillers for biological filters are useful in aquaculture. The reaction kinetics study is the next step after confirming the suitability of the filling used. The operation of modern aquaculture systems and their further dynamic growth requires the continuous development of new technologies [24]. It is equally important to keep in mind while developing new technologies or solutions that biological filter operation is affected by a number of factors, e.g., dissolved oxygen and the rate of its diffusion into the biomembrane, temperature, pH, salinity and organic matter burden on the biological filter, which is not always reflected in laboratory tests [7,25–27]. Considering the above, a two-step experiment was conducted on a semi-commercial scale, in which submerged and dripping biological filters with three different fillings were used, which are different from each other. CFs (commercial fitting HXF12KLL) have been designed to obtain the largest possible area in a unit of volume. In addition, their spatial structure allows the free flow of water through the centre of individual fittings, which effectively supports nitrification (nitrifiers have constant access to nitrogen compounds). PSCs (polyethylene screw caps for PET bottles) have more than three times smaller surface area per unit volume, but due to their shape, the inflow of nitrogen compounds to the biofilm is greatly facilitated. PPAs (polypropylene aggregate) have the least regular shape. A microscopic photograph reveals a lack of internal spaces on which a bacterial biofilm could develop. However, the outer structure has numerous recesses and protrusions, which are very diverse

and provide favourable conditions for the development of nitrifiers. The main donor of nitrogen compounds were common carp *Cyprinus carpio* juveniles. The study examined the dynamics of nitrogen compound transformations during biological bed maturation and matured bed operation. Moreover, increasing the daily nutrient dose allowed the bed operation to be examined with increasing loads.

A system with submerged filters with commercial fillings was used as a control. During the experiment, the hypothesis was verified that the use of polypropylene aggregate and polyethylene screw caps for PET bottles as fillers for biological filters will achieve the same effects as the use of commercial fittings HXF12KLL.

### **2. Materials and Methods**

The experiment lasted 60 days in the submerged biological bed variant (two days of fish acclimatisation and 58 days of the water parameter measurements) using tap water. The experiment lasted 45 days in the dripping biological bed variant (two days of fish acclimatisation and 43 days of the water parameter measurements). The biological bed maturated during this time, which allowed the dynamics of nitrogen compound transformations in a maturated bed system to be examined. Moreover, increasing the daily nutrient dose allowed the bed operation to be examined with an increasing load. Ammonium, nitrites and nitrates' concentration in tap water at the beginning of the experiment was 0.045 mg N-NH4/dm3, 0.027 mg N-NO2/dm<sup>3</sup> and 0.872 mg N-NO3/dm3.

### *2.1. Experimental Conditions*

### 2.1.1. RAS Systems

The experiment was conducted in four identical semi-open RAS systems with 625 dm<sup>3</sup> volumes each (daily refill—150 dm3) modelled on devices used by Sikora et al. [18]. Three types of biological filter fillings were examined:


**Figure 1.** Polypropylene aggregate (PPA)-specifics surface area—ca. 12 m<sup>2</sup> (**A**); polyethylene screw caps for PET bottles (PSC)-specifics surface area—ca. 9 m2 (**B**); Commercial fittings HXF12KLL (CF)-specifics surface area—ca. 25.5 m2 (**C**).

Each RAS system consisted of: an upper retention reservoir, a two-chamber lower retention reservoir (sedimentation chamber, pump chamber) and two rearing tanks. The systems were also fitted out with heaters, thermostats and aeration systems. The biological filter filling volume was 30 dm3. The specifics surface area of the biological filters used was: PPA—ca. 12 m2, PSC—ca. 9 m2, CF—ca. 25.5 m2. For submerged filters (Figure 2), the filling was placed directly in the sedimentation chamber. The filling was suspended above the sedimentation chamber in the variant with dripping filter (Figure 3), which had sprinklers mounted above them.

**Figure 2.** Recirculation aquaculture system with submerged filter scheme: 1 upper retention tank—0.251 m3, 2 rearing tank—2 <sup>×</sup> 0.096 m3, 3 lower retention tank (3a chamber with biofilter, 3b chamber with pump)—0.182 m3, 4 water inlets, 5 sprinklers, 6 water outlets, 7 pump, 8 water supply to upper retention tank, 9 submerged filter—0.030 m3.

**Figure 3.** Recirculation aquaculture system with dripping filter scheme: 1 upper retention tank—0.251 m3, 2 rearing tank—2 <sup>×</sup> 0.096 m3, 3 lower retention tank (3a chamber with biofilter, 3b chamber with pump)—0.182 m3, 4 water inlets, 5 sprinklers, 6 water outlets, 7 pump, 8 water supply to upper retention tank, 9 dripping filter—0.030 m3.

The systems were started five days before the fish were put in them to stabilise the conditions in the recirculation systems. The circulating water temperature was 25.0 ± 0.1 ◦C. Before the experiment, the systems were first thoroughly dried, disinfected with potassium permanganate at 7.5 g per run and then rinsed with water from the water supply system for 24 h.

### 2.1.2. Fish

Common carp juvenile (*Cyprinus carpio*), bred at the Department of Lake and River Fisheries of the University of Warmia and Mazury in Olsztyn and cultured under controlled conditions, were used in all experiment as a source of nitrogen compounds. Out-of-season carp breeding was carried out in accordance with the methodology described by Kucharczyk et al. [28], with the insemination method modified by Kucharczyk et al. [29] for burbot (*Lota lota*), ide (*Leuciscus idus*) and asp (*Aspius aspius*). The carp larvae were reared in a recirculation system at 25 ± 0.1 ◦C and fed brine shrimp (*Artemia salina*) nauplii for the first 25 days. Artificial feed with a granulation of 0.5–1.1 (Skretting, Norway: raw protein content 54%, fat content 18%) was introduced after that time.

In the submerged biological bed variant, 1379 ± 49 g (average ± SD) of fry with a weight and unit length of 3.76 ± 1.55 g (average ± SD) and 59.00 ± 8.77 mm (average ± SD) were placed in the RAS systems. In the dripping biological bed variant, however, 1400.00 ± 0.00 g fry with a unit weight and length of 16.30 ± 4.80 g and 96.40 ± 9.65 mm were placed in the RAS systems. All fish used during the experiment are juvenile forms of common carp.

The fish (carp) were fed twice daily (at 8.30 and 15.00) during the experiment with an artificial feed with a granulation of 1.1 (Skretting, Norway: raw protein content 54%, fat content 18%) and 1.9 mm (Skretting, Norway: raw protein content 50%, fat content 20%), with a single feeding lasting 30 min. The feeding dose was set at 3% of the daily biomass (half of the daily feeding dose was provided in one feeding). Subsequently, the dose was increased daily by 3% of the initial fish biomass [30,31]. The initial daily feed dose was 20.9 ± 0.5 g (average ± SD).

Sodium chloride was added to circulating water beginning on the day the nitrite nitrogen concentration reached 1 mg N-NO2/dm3. Nitrite shows of affinity to the Cl- /HCO3 - ion exchange. As a result, a part of the nitrite is taken up by fish instead of chloride. Increase of chloride concentration in the water reduces nitrite uptake by fish [32].

### *2.2. Water Quality Measurements*

Water tests started two days after the fish were placed in the systems. The water in each system was examined at 11:00 (samples were taken from the sprinkler). Hach Lange cuvette tests with a dedicated DR 5000 spectrophotometer were used for the water testing [18]:

Ammonium—cuvette test LCK 304 Nitrite—cuvette test LCK 341 Nitrate—cuvette test LCK 340

Daily ammonium assays were conducted for the initial 17 days of the experiment. The aim of this test was to check whether the compound accumulates in the water in excessive amounts. After that time, measurements were conducted every three days to monitor the total ammonium concentration in the circulating water. The nitrite concentration was checked daily throughout the experiment to determine the time needed to start stage I of nitrification (oxidation of ammonia to nitrites). The nitrate concentration in the water was tested daily for the first 28 days, after which measurements were conducted every two days to determine the time needed to start stage II of nitrification (oxidation of nitrite nitrogen to nitrate nitrogen).

Moreover, the following water parameters in the bed were measured at 11.00:


Moreover, the following water parameters were tested before the morning and afternoon feeding:


### *2.3. Fish Measurements*

The fish were measured (average weight and average total length) at the beginning and end of the experiment. Thirty fish were collected randomly and weighed with an analytical balance (KERN & Sohn GmbH, Balingen, Germany) with an accuracy of 0.1 mg. The fish body lengths were measured with a calliper MEGA 20513 (Profix, Warsaw, Poland) with an accuracy of 0.01 mm. Due to the fish size, their bodies were measured after the experiment with a ruler with an accuracy of 1 mm. Moreover, the fish biomass was determined at both the beginning and end of the experiment (analytical balance KERN & Sohn GmbH, Germany). The fish were anaesthetised during the measurements with an MS222 anaesthetic at a concentration of 50 ppm.

### *2.4. Statistical Analysis*

The dynamics of nitrogen compound (ammonia, nitrites, nitrates) transformations were examined with a Kruskal–Wallis ANOVA test on ranks (*p* > 0.05) and subsequently by multiple comparisons of mean ranks (*p* > 0.05) for all samples. The distribution normality was verified with the Shapiro–Wilk test (*p* > 0.05) before the ANOVA test was performed. All results were analysed statistically using Statistica 13.1 software (StatSoft, Tulsa, Oklahoma, USA). In addition, in order to analyse the nature of individual stages of nitrification, regression equations were performed and the correlation coefficient (R) was calculated.

### **3. Results**

### *3.1. Variant I—Use of Submerged Biological Filter*

### 3.1.1. Ammonium Nitrogen

At the beginning of the experiment, the ammonium concentration increased rapidly in all experimental RAS and was about 3 mg N-NH4/dm3 from day 3 (Figure 4).

**Figure 4.** Ammonium nitrogen concentration in experimental systems using submerged biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL.

High ammonia levels lasted for several days and then quickly decreased. After this period, until the end of the experiment, no major increases in ammonium were recorded in RAS circuits (Table 1).

**Table 1.** Ammonium nitrogen concentration in RAS, in which various tested fillings (PPA polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL) were used for two separated stages, Stage I (high ammonium concentration) and Stage II (low concentrations ammonium) separated by a period of rapid drop in ammonium (break down).


### 3.1.2. Nitrite Nitrogen

In all RAS, nitrite levels increased for several dozen days until the maximum concentration was reached. After this time, the concentrations decreased until they reached a relatively low and stable level, which remained until the end of the experiment (Figure 5; Table 2).

**Figure 5.** Nitrite nitrogen concentration in experimental systems using submerged biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL.

**Table 2.** Nitrite nitrogen concentration in RAS, in which various tested fillings (PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL) were used for two separated stages, Stage I (high nitrite concentration) and Stage II (low nitrite concentration).


### 3.1.3. Nitrate Nitrogen

The nitrate concentration in the system with PPA as the biological filter filling increased until day 53 and reached 32.05 mg N-NO3/dm3. It began to decrease afterwards. A similar trend was observed in the other two experiment variants. The concentration in the PSC system also rose until day 53 to reach a similar level (34.35 mg N-NO3/dm3). The highest concentration in the system with CF filling was higher than in the other two systems (50.40 mg N-NO3/dm3) and it was measured during a shorter time (day 44) (Figure 6).

**Figure 6.** Nitrate nitrogen concentration in experimental systems using submerged biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL.

### 3.1.4. Other Water Parameters

The average water temperature in the system with PPA was 24.8 ± 0.4 ◦C, average pH—8.01 ± 0.15, whereas the average concentration of oxygen dissolved in water and saturation were 6.6±0.7 mg O2/dm3 and 79.8 ± 8.7%, respectively. The same parameters in the PSC system were: water temperature 24.8 <sup>±</sup> 0.3 ◦C, pH 8.03 <sup>±</sup> 0.15, dissolved oxygen concentration 6.7 <sup>±</sup> 0.7 mg O2/dm3, saturation 82.4 ± 8.6%. In the system with CF filling, the parameters were: water temperature 24.8 ± 0.6 ◦C, pH 8.06 <sup>±</sup> 0.14, dissolved oxygen concentration 6.8 <sup>±</sup> 0.6 mg O2/dm3 and saturation 81.6 <sup>±</sup> 7.2%.

### *3.2. Variant II—Use of Dripping Filters*

The system with a submerged filter with CF filling (CFsf) was used again as the control in the second part of the experiment with dripping filters. There was a failure and leak in the system with the dripping filter filled with CF during this part of the experiment. The system was restarted in accordance with the procedure presented in the methodology.

### 3.2.1. Ammonium Nitrogen

As in the case of submerged filters, the concentration of ammonium in the experimental RAS systems increased rapidly and from day 3 it was about 3 mg N-NH4/dm3 (Figure 7).

**Figure 7.** Ammonium nitrogen concentration in experimental systems using dripping biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL.

High ammonia levels lasted for several days and then quickly decreased. After this period, no significant increases in RAS ammonium were noted until the end of the experiment (Table 3).

**Table 3.** Ammonium nitrogen concentration in RAS, in which test fillings (PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL) were used for two separated stages: Stage I (high ammonium concentration) and Stage II (low ammonium concentrations) separated by a period of rapid drop in ammonium concentration (break down).


Only in the system in which a failure occurred were other characteristics of ammonium concentrations noted. The highest concentration in the system with CF was recorded on day 3—1.230 mg N-NH4/dm3—and it decreased afterwards. The ammonium concentration decreased below 0.3 mg N-NH4/dm3 after day 8 and until the end of the experiment it fluctuated, with an average concentration of 0.188 <sup>±</sup> 0.047 mg N-NH4/dm3 (Figure 7).

### 3.2.2. Nitrite Nitrogen

In all RAS, nitrite levels increased for several dozen days, until the maximum concentration was reached. After this time, the concentrations decreased until they reached a relatively low and stable level, which remained until the end of the experiment (Figure 8; Table 4).

**Figure 8.** Nitrite nitrogen concentration in experimental systems using dripping biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL.

**Table 4.** Nitrite nitrogen concentration in RAS, in which test fillings (PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL) were used for two separated stages: Stage I ((high nitrite concentration) and Stage II (low nitrite concentration).


### 3.2.3. Nitrate Nitrogen

The nitrate concentration in the system with PPA as the biological filter filling increased until day 44 and reached 32.150 mg N-NO3/dm3. It decreased slightly afterwards. A similar trend was observed in the system with PSC. The highest concentration of 30.00 mg N-NO3/dm<sup>3</sup> was observed on day 40 and it decreased slightly afterwards. A different situation was observed in the system with CF. The concentration of nitrate nitrogen increased to 29.800 mg N-NO3/dm3 on day 36; afterwards it decreased until the end of the experiment, reaching 25.00 mg N-NO3/dm<sup>3</sup> on the last day. In the system with CFsf, the nitrate concentration increased until day 32 and reached 32.15 mg N-NO3/dm3 and it decreased slightly afterwards (Figure 9).

**Figure 9.** Nitrate nitrogen concentration in experimental systems using dripping biological filters: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL.

### 3.2.4. Other Water Parameters

The average water temperature in the system with PPA was 24.9 ± 0.1 ◦C, average pH—8.84 ± 0.10, whereas the average concentration of oxygen dissolved in water and saturation was 7.3 ± 0.3 mg O2/dm3 and 88.9 <sup>±</sup> 3.8%, respectively. The same parameters in the system with PSC were: water temperature 24.9 <sup>±</sup> 0.2 ◦C, pH 8.82 <sup>±</sup> 0.10, dissolved oxygen concentration 7.5 <sup>±</sup> 0.4 mg O2/dm3, saturation 92.1 ± 3.9%. In the system with CF, the parameters were: water temperature 25.0 ± 0.3 ◦C, pH 8.83 <sup>±</sup> 0.12, dissolved oxygen concentration 7.0 <sup>±</sup> 0.4 mg O2/dm<sup>3</sup> and saturation 85.7 <sup>±</sup> 4.7%. The parameters in the control system (CFsf) were: water temperature 24.9 ± 0.1 ◦C, pH 8.74 ± 0.15, dissolved oxygen concentration 7.0 <sup>±</sup> 0.3 mg O2/dm3 and saturation 86.1 <sup>±</sup> 2.7%.

### *3.3. Statistical Analysis of Results*

The results do not have a normal distribution (*p* < 0.05, Shapiro–Wilk test). A statistical analysis of the results using a Kruskal–Wallis ANOVA rank test revealed significant differences between the RAS systems (*p* < 0.05). Detailed results are provided in the table (Table 5).

**Table 5.** Analysis of concentrations of nitrogen compounds using Kruskal-Wallis ANOVA test on ranks (*p* > 0.05) showed differences between individual RAS systems. Results of a multiple comparison test of mean ranks for all samples: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL. Data marked with the same letter in rows did not differ statistically.


Analysis of the results of ammonium concentrations in individual RAS allowed the determination of two periods of a different character for which regression equations take a different form. After placing the fish, the concentration of ammonium in individual systems increased. High levels of ammonium were noted from day 3, followed by low ones. There was one day of intermediate values between periods. Increases and decreases of ammonium concentration occurred rapidly. In the case of a system in which a failure occurred and it became necessary to restart it instead of a period with high concentrations, there was a period during which the concentration gradually decreased (Tables 1 and 3).

The analysis of nitrite concentration resulting in individual RAS allowed the determination of four periods with a different nature of the course. These periods for different RAS systems have a different duration but show a similar nature. In the first period, the regression equation describing the increase in concentration takes the form of an exponential function. In the second period, the concentration initially rose, reached the maximum level and began to decline and the equation takes a polynomial form. In the third period, characterized by a decrease in concentration, the equation takes on a power form. In the fourth period, nitrite concentrations remain at the same level without significant decreases or increases, and the equation again takes on a polynomial form (Table 6a,b).

**Table 6.** (**a**) Nitrite nitrogen concentration analysis in experimental RAS systems using regress equation (Reg.) and correlation coefficients (R): PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL. (**b**) Nitrite nitrogen concentration analysis in experimental RAS systems (dripping, CF) using regress equation (Reg.) and correlation coefficient (R).


The concentration of nitrates in RAS systems show a similar tendency. After the initial slow increase in concentration, accelerated growth followed and was then inhibited. An analysis of the obtained results showed a decrease in nitrate concentration at the end of the experiment, which is reflected in the derived regression equations (Table 7).



### *3.4. Fish*

During the experiment, no mortality was observed in the reared fish.

### **4. Discussion**

In the experiment, the dynamics of changes in nitrogen compounds in the water used for rearing fish in semi-closed RAS were analysed in detail. The dynamics of nitrogen compounds (ammonium, nitrite, nitrate) during the maturation of biological filters as well as during the operation of mature filters were examined. Furthermore, the usefulness of PPA and PSC as fillings for biological filters was demonstrated. Due to the differences in construction, all three fillings have their own advantages and disadvantages. Therefore, when comparing fillings, the analysis was based on the analysis of changes in concentrations of nitrogen compounds in water and the time needed to achieve full filter functionality. Comparing structural parameters, especially similar in terms of the size of the active surface of PSC and PPA fillings, could be subjective. The impact of structural parameters should be examined at a later stage of the study, in which the impact of individual factors (e.g., oxygen, temperature, suspense solids, diffusion rate, etc.) would be analysed on the dynamics of the nitrification process. Water exchange is necessary in aquaculture farms using a closed water cycle equipped with biological filters without denitrification due to ever increasing level of nitrate. Denitrification is the process of reducing nitrates to gaseous nitrogen. It is an anaerobic process that can sometimes occur in biological filters overloaded by biogens. The amount of water needed varies depending on several factors. The most important is the increase in nitrate concentration, which in high concentrations shows adverse effects on aquatic organisms [21,23,33]. At the start-up stage of RAS facilities, water change is required for a similar reason. High concentrations of toxic forms of nitrogen occurring during the maturation of biological filters require the use of water changes not only to supplement evaporation. In the experiment, a one-time top-up of 150 dm3 daily was used. This allowed for the dilution of nitrogen compounds and while supplementing calcium compounds in water, it also allowed better observation of the dynamics of nitrogen compound transformation processes. An important factor was also the use of common carp—a model species which is considered to be relatively resistant to adverse environmental rearing conditions [34].

Mechanical and biological filters must work together for the proper functioning of aquatic organisms [7,26,27,35,36]. The importance of the filtration process becomes significant when the breeding facility is built in RAS technology [22,37–42]. For the proper course of biological purificationin the nitrification process, the cooperation of two groups of microorganisms is required. The first group oxidizes ammonia to nitrites and the other group oxidizes the formed nitrites to nitrates [6,7,43–45]. By analysing the concentration of particular forms of nitrogen in the water used during the commissioning of breeding facilities, it is possible to observe individual stages of nitrification and thus determine whether the biological filter has gained functionality [18,45]. The unpredictability of the maturation process was shown in an experiment carried out by Pulkkinen et al. [46] illustrating the effect of the type of biological filter used (fixed and moving bed bioreactors) on nitrification in recirculating aquaculture systems. In that experiment, despite the use of filter fillings operating for six months, the faulty operation of the bed was observed, which was revealed by an initially high concentration of nitrites, whose low level was achieved only after about eleven weeks. According to these authors, such a situation was caused by disturbances in the bacterial composition produced by the shock of transportation to a laboratory. This indicates the exceptional sensibility of nitrifying bacteria to variable environmental conditions. Due to the high concentration of toxic nitrogen compounds, fish rearing in circuits with immature biological filters threatens the loss of fish being reared. This is especially true for salmonids [47]. Therefore, species resistant to elevated concentrations of nitrogen compounds should be used at the start-up stage of aquaculture facilities.

The experiment using fish as a source of compounds needed in the nitrification process allowed observation of the work of the biological filter in conditions simulating real fish breeding [18]. Research carried out in this way is very important to properly carry out the maturation process of the biological filter and obtain a filter adapted to the given conditions in RAS systems. A review of the literature has shown that this approach to the topic is rare [18,48]. Research using synthetic solutions with nitrogen compounds dominates the field [40,49,50]. In this experiment, carp was used as a model species [34] with high metabolism [31].

During the experiment, all systems noted a rapid increase in the ammonium content in water to approximately 3 mg N-NH4/dm<sup>3</sup> (Figures 4 and 7; Tables 1 and 3). The increase in ammonium concentration resulted from placing the fish in the water cycle and starting feeding them. It should be emphasized that carp is a species with very high metabolism [31]. This level persisted from 4 days (submerged filter, CF) to 9 days (submerged filter, PSC) (Figure 4). In other filters, the duration of elevated ammonium concentration in the circuits was 7–8 days. After this period, there was a rapid, several-fold decrease in ammonium concentration in the RAS. It was assumed that the procedure involving the drying of the circuits and their disinfection through the use of potassium permanganate will eliminate nitrifying bacteria. The abrupt increase in ammonium has proved this assumption. Low concentrations persisted until the end of the experiment. In the system that was restarted, the distribution of ammonium concentrations was different (Figure 7). The first recorded results were about 60% lower than in other systems, and there was a rapid decrease in ammonium concentration. The fall curve was much milder. There is a discrepancy between the results obtained for two systems in which submerged filters filled with CF were used (Figures 4 and 7). The duration of elevated ammonium concentration for these systems was four days and eight days. The results during the experiment are divergent from the experiment carried out by Sikora et al. [18], in which no increase in ammonia concentration was observed. Kuhn et al. [45] reported similar results. In their research, they compared the effects of biological filters that were inoculated with nitrifying bacterial cultures with uninoculated filters. In unvaccinated RAS systems, ammonium nitrogen concentrations quickly rose above 2.5 mg/dm<sup>3</sup> and then decreased. The time in which the ammonium nitrogen concentration was increasing and decreasing differed for the studied filters. The curves of changes in concentration were also different. In the case of filters inoculated with bacterial cultures, the course was similar to that observed in the system, which was restarted. This example supports the suggestion that nitrifying bacterial cultures remain in this system despite disinfection with potassium permanganate. For studies based on nitrifiers immobilized in PVA (polyvinyl alcohol) and then adapted to the salty environment [51], the time needed to remove the elevated TAN concentration (initial concentration of TAN introduced into bioreactors was 10 mg/dm3) to nitrite ranged from 22 days (salinity 30 ppt)

to 26 days (salinity 7.5 ppt). These are longer periods than obtained during the author's experiment. In 2015, Hu et al. [52] conducted research on the removal of nitrogen compounds from tilapia culture (*Oreochromis niloticus*) using aquaponic crops (tomato *Lycopersicon esculentum* and pak choi *Brassica campestris L. subsp. chinensis*). In these studies, they used root systems of cultured plants as growth surfaces for nitrifying bacteria. During the study, the highest TAN concentrations were reached when the TAN concentrations peaked around day 7 and were around 25 mg/dm<sup>3</sup> (tomato) and 32.5 mg/dm<sup>3</sup> (pak choi). These concentrations are higher than those observed in the discussed experiment; moreover, they did not decrease to a similar degree in a comparable period of time. This difference is probably due to the smaller root surface area in relation to the effective surface of the applied biological filter medias.

According to Karpinski et al. [53], the increase in nitrite concentration is delayed in relation to the increase in ammonium concentration (Figures 4, 5, 7 and 8; Tables 1, 3 and 6a). In the conducted experiment, an increase in nitrite concentrations was observed from the beginning of the experiment (Figures 5 and 8). Initially, these concentrations were relatively low compared to the observed concentrations of ammonium nitrogen, however, since the growth curve is exponential, the concentration of nitrite increased very quickly (Table 6a). Subsequently, the increase in nitrite concentration slowed down and slightly collapsed, followed by a reversal of the trend observed in the first period. Nitrite concentrations in all RAS systems dropped rapidly, reaching a relatively constant level (Figures 5 and 8, Table 6a,b). This course of concentration changes was observed in all RAS systems except the system in which the failure occurred (Table 6a). In this system, the period of deceleration and collapse of the upward trend does not occur (Table 6b). Maximum concentration values were reached much faster and they began to decrease faster. This is reflected in the equations describing the course of nitrification (Table 6a,b). The time needed to stabilize the second phase of nitrification ranged from 15 to 33 days (Figures 5 and 8). This is different from the results obtained by Kuhn et al. [45]. Despite the fact that measurements were carried out for 28 days, no downward trend was observed. This situation occurred in systems not inoculated with nitrifying bacteria. Different results were obtained in systems inoculated with nitrifiers—no increase in nitrite concentration was observed. In the research conducted by Sikora et al. [18], the time needed to stabilize the second phase of nitrification was 35 days, which is longer than the results observed in this article, although similar. Comparable times to those obtained during the experiment and to those obtained by Sikora et al. [18] were also obtained by Seo et al. [51] during an experiment with the acclimatization of nitrifiers to saltwater conditions. The acclimatization time of biological filters, and thus the time needed for nitrite oxidation to a safe level, was achieved after 33 days (salinity 7.5 and 15 ppt) and 39 days (salinity 30 ppt). Longer nitrite elimination times than observed during the experiment were also reported by Hu et al. [52]. The time needed for oxidation of nitrite to nitrate in aquaponic cultivation was about 40 and 50 days for tomato and pak choi respectively. As in the case of ammonia oxidation, the surface of the roots on which nitrifying bacteria developed was smaller than in the tested biological filter medias.

For nitrates, the recorded concentration values increased from the very beginning of the experiment. Initially, as with nitrites, this increase was slow to eventually accelerate. After a period of dynamic growth, nitrate concentrations stabilized (Figures 6 and 9; Table 7). For CF, despite the similar nature of the increase curve, the concentrations were higher during the first part of the experiment. No similar trend was observed for the same type of filter in the second part of the experiment (Figures 6 and 9). Higher values in the initial phase of growth in the second part of the experiment show the concentrations in the RAS system in which the failure occurred (Figure 9). The similar nature of the increase in nitrate concentration was reported by both Kuhn et al. [45] (circulation vaccinated with nitrifying bacteria) and Sikora et al. [18]. Kuhn et al. [45] did not observe an increase in the concentration of nitrates in the RAS system inoculated with nitrification, which was associated with the incomplete nitrification process. The nitrate concentration increase curve presented by Seo et al. [51] was comparable to the present experiment. At the end of the experiment, a decrease in nitrate concentration was observed in individual RAS systems. Probably this decrease was associated with the development of heterotrophic bacteria in the volume of water.

During this experiment, it was analysed how the concentrations of nitrogen forms in water are shaped in experimental RAS systems. On this basis, the course of the nitrification process was determined along with the time needed for the biological filter to mature. The maturation of the biological filter is largely dependent on the temperature at which the process takes place (Table 8). In systems using cool water, a mature filter can be obtained after a few months of its work, while this period is significantly shorter in systems developed for the needs of thermophilic species. Another factor that may affect the maturation of biological filters is the presence of the desired nitrifying bacteria. Biological filters inoculated with nitrifying bacteria, operating at 25 ◦C, show desirable properties after 53 days [54]. The use of water from tilapia culture (temperature 26.4 ◦C) allowed a functional biological filter to be obtained after 56 days [5]. These results do not reflect the information provided by Kolman [55], who states that at 18 ◦C, it takes 40 to 60 days to obtain a working biological filter. In addition, the periods given differ from those obtained by Sikora et al. [18]. Studies have shown that the time needed to mature a biological filter without inoculating the culture cycle with nitrifying bacteria cultures is 35 days at 23 ◦C.

**Table 8.** Comparison of time needed to mature biological filter in accordance with temperature on the example of literature data and conducted experiment: PPA—polypropylene aggregate, PSC—polyethylene screw caps for PET bottles, CF—commercial fittings HXF12KLL, CFsf—control submerged filter for dripping filters, commercial fittings HXF12KLL.


The conducted tests showed that at 25 ◦C, without introducing nitrifying bacteria cultures into the RAS system, the time needed for the filter to mature is from 21 to 33 days. If nitrifying bacteria are present, this period may be shorter. In one of the RAS systems used for the experiment, a failure occurred. As a result of the failure, it was necessary to restart the RAS. This RAS was dried and disinfected in accordance with the adopted procedure. In this system (drip filter, CF), a functional biological filter was obtained on day 16.

### **5. Conclusions**

The maturation of biological filters is, despite the overall repetitive pattern, a largely variable process. The time of individual stages of nitrification and their courses differ. It largely depends on the type of filter used and its filling and prevailing conditions. The obtained results showed the effectiveness of all tested biofilter media. The shortest time to biofilter maturation was when submerged CF was used. However, the maximum peak concentration of nitrite was also noted when it was used. For fish culture welfare, it is important that nitrite concentration should be as low as possible. Therefore, other tested biofilter media worked better from this point of view. New technologies are being sought, new materials are used as filter fillings and nitrification conditions are being modified. Increasingly, breeders want to increased production faster and faster. This involves the use of increasingly efficient filters that can be run faster. Research such as that described in this article is helpful, even necessary,

in learning how the nitrification process works, the conditions that shape it and what should be done to obtain fully efficient biological filters.

In addition, the article describes an unorthodox approach to new materials useful as cartridges for biological filters. Both PPA and PSC are materials that are associated as a convenient intermediate, not a useful cartridge for biological filters. However, these materials proved to be effective within the assumed parameters.

**Author Contributions:** The following statements should be used "Conceptualization, M.S., D.K.; Methodology, M.S., J.N.; Software, M.S.; Validation, M.S., J.N.; Formal Analysis, M.S.; Investigation, M.S.; Resources, M.S., D.K., J.N.; Data Curation, M.S.; Writing-Original Draft Preparation, M.S.; Writing-Review & Editing, M.S., J.N., D.K.; Visualization, M.S.; Supervision, D.K.; Project Administration, D.K. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by UWM Olsztyn, project No. 18.610.005-110 and the project was financially supported by Minister of Science and Higher Education in the range of the program entitled "Regional Initiative of Excellence" for the years 2019-2022, Project No. 010/RID/2018/19, amount of funding 12.000.000 PLN.

**Conflicts of Interest:** The authors declare no conflict of interest.

### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

### *Article* **Is It Possible to Restore a Heavily Polluted, Shallow, Urban Lake?**

### **Jolanta Grochowska \*, Renata Augustyniak, Michał Łopata and Renata Tandyrak**

Department of Water Protection Engineering and Environmental Microbiology, Institute of Engineering and Environmental Protection, University of Warmia and Mazury, St. Prawoche ´nskiego 1, 10-720 Olsztyn, Poland; rbrzoza@uwm.edu.pl (R.A.); michal.lopata@uwm.edu.pl (M.Ł.); renatat@uwm.edu.pl (R.T.) **\*** Correspondence: jgroch@uwm.edu.pl

Received: 17 April 2020; Accepted: 25 May 2020; Published: 27 May 2020

**Abstract:** The research was carried out on Karczemne Lake, a water reservoir located in Kartuzy (northern Poland, Pomeranian Lake District). Monitoring of the water and bottom sediment of Karczemne Lake showed a very high level of contamination of the reservoir by a long-term inflow of untreated municipal sewage. The trophic status index of total phosphorus (TP) was unusually high at 101, and the TP content in the bottom sediments—31 mg g−<sup>1</sup> (dry weight)—was the highest value recorded worldwide in a lake. Based on the monitoring results, to achieve constant improvement of the water quality, we recommend a completely new, safe and economically justified method of bottom sediment removal and management. A very important aspect of this method is the prevention of uncontrolled sewage discharge back into the lake basin. Removed sediment with interstitial water will be pumped through a pipeline and transported to a sewage treatment plant. In the sediment mining field in which the sludge will be removed, the first phase of phosphorus inactivation will be carried out to chemically precipitate pollutants distributed in the water column as a result of sediment resuspension. After the deepening of the entire lake basin, the method of phosphorus inactivation will be carried out on the entire surface of the lake as the next stage of restoration. A supporting activity will be biomanipulation. Before the restoration is started, the municipal sewerage system will be modernized.

**Keywords:** urban lake; restoration; dredging; phosphorus inactivation; biomanipulation

### **1. Introduction**

Strong anthropopressure in the catchment areas of lakes (urbanization, industrialization, deforestation and intensive livestock), which stimulates an increased supply of nutrients to waters, causes the acceleration of eutrophication [1–3]. Accelerated eutrophication is exemplified by strong cyanobacterial blooms, species depletion at all trophic levels, the disappearance of valuable fish species and the deterioration of the taste and smell of the water, which leads to the cessation of the reservoir being used for municipal and recreational purposes [4–6]. A sign of excessive eutrophy of the lake is an increase in nutrient concentrations (nitrogen and phosphorus) in the water and an uneven distribution of the amount of oxygen in the water column [7–9]. The surface layers of the reservoir are usually saturated with oxygen, while at the bottom, anaerobic conditions prevail, promoting the release of nutrients from bottom sediments. In oligotrophic and mesotrophic lakes, bottom sediments are a "trap" that bind excess nutrients, especially phosphorus. In degraded lakes, in which the bottom water layers are deoxygenated and therefore accompanied by a reduction in redox potential, the reverse process takes place. The stored substances, such as nutrients, are released into the near-bottom water layers [10–12]. This phenomenon is called internal loading. The amount of nutrients in the bottom sediment of the lake is extremely high. The sediments of degraded lakes are the main, inexhaustible source of biogenic compounds. Considering the content of the most important

nutrient—phosphorus—in the particular parts of lake ecosystems, it was found that approximately 90% of the total phosphorus was contained in the upper 10-cm layer of the bottom sediments [13]. This result clearly indicates the need to limit this "source". The inhibition of internal loading is the main goal of all lake restoration methods, including artificial aeration, phosphorus inactivation, capping and the Ripl method [14–16].

The removal of bottom sediment is widely regarded as a radical but highly effective restoration method for shallow, heavily degraded reservoirs [17]. The complete removal of bottom sediment to the parent rock floor guarantees radical renovation of the lake and an increase in its volume, which often has previously been decreased due to the inflow of excessive pollution. Considering the thickness of sediments in postglacial lakes, often up to several meters or more, this method should be considered unrealistic.

Innovative lake restoration projects using the dredge method should be preceded by detailed monitoring of water and bottom sediment. Therefore, lake monitoring should be started with an examination of the spatial composition of bottom sediments to precisely determine the thickness of the most contaminated layer that is supposed to be removed. In addition, uncovered deposits that remain in the lake should be poor in nutrients, while phosphorus, as the main element responsible for eutrophication, must be stored in a biologically inactive form, i.e., as residual phosphorus or calcium-bound phosphorus [18–20]. This method of removing bottom deposits is very rarely used due to its complexity, cost and implementation difficulties. However, although the name of the method refers to bottom sediment removal, such a project must also take into account the output development and processing/utilization of the solid and liquid fractions [15,16,21]. The processing, management and utilization of deposits is often the greatest logistical and economic challenge.

An example of a very inept attempt at lake restoration by the bottom sediment removal method was the case of Mogile ´nskie Lake in Poland. The sediment and interstitial water removed from the lake were not properly managed, and before they were processed, the effluents returned to the lake water, destroying the initial effects and causing an ecological catastrophe [22].

In the case of the restoration of the Swedish Lake Trummen, the sediment was pumped to simple settling ponds constructed in an abandoned farming area from which the topsoil had first been removed. The runoff water from the settling ponds—a mixture of lake and interstitial water—was treated with aluminum sulfate in a simple plant for the precipitation of phosphate and suspended matter. Before restoration, the total phosphorus content of the lake water was approximately 600 μg L<sup>−</sup>1. The phosphorus concentration of the water from the settling ponds was on the order of milligrams per liter before the treatment. After precipitation, the total phosphorus content of the runoff water was approximately 30 μg/L [23]. The area designated for the infrastructure to process the liquid and solid output was approximately 30 ha. It should be noted that it is rare to find so much space in the vicinity of a lake that can be used to safely convert the removed deposit.

According to a study of the Vajgar artificial reservoir by Björk et al. [24], the top of the sediment that had accumulated in the pond served as an uncontrolled source of nutrients and was removed. The sediment pumping started in August 1991, and by the end of 1992, approximately 330,000 m<sup>3</sup> of sediment had been pumped out of the pond and transported in pipes to seven settling lagoons approximately 2.5 km away, each of which was approximately 1 ha in area and 3–5 m deep. The sediment transported from the fish pond had a dry mass content of 10–15%. The whole cost was approximately US \$850,000.

The above examples indicate that the removal of bottom sediments from reservoirs is logistically very difficult, and in the case of ill-considered solutions, it can cause the opposite effect—an ecological disaster. Another disadvantage of this method is the very high cost; therefore, before developing a project to extract sediments, the spatial distribution of the phosphorus in the sediment should be determined to accurately indicate the thickness of the sediment, the extraction of which will guarantee the improvement of the state of the lake as a result of stopping the emission of pollutants from the sediments back into the water column.

The aim of the study was to determine the spatial distribution of pollutants in bottom sediment and interstitial water of Karczemne Lake and on their basis to determine the thickness and volume of sediment necessary for extraction to improve the quality of water in the reservoir. This research has allowed the development of the concept of a completely new, safe and economically justified method of mining and managing the bottom sediments. The proposed method will prevent uncontrolled effluent drainage back into the lake basin. In addition, it does not require the construction of a completely new bottom sediment treatment system consisting of presses, centrifuges, polymer-dosing stations, water-conditioning equipment and reaction pools. After sediment is removed from the lake bowl, phosphorus inactivation and biomanipulation is also recommended.

### **2. Material and Methods**

### *2.1. Study Site*

Karczemne Lake (54◦19 42" N, 18◦11 27" E) is a strongly degraded urban lake located in the Kaszubian Lake District and belongs to the macroregion of the Eastern Pomeranian District in Kartuzy town [25]. Karczemne Lake is a shallow, polymictic, flow-through lake. Its area is 40.4 ha, and its maximum depth is 2.3 m (Figure 1). More detailed morphometric parameters of the lake are given in Table 1.

**Figure 1.** Location of the Karczemne Lake.


From the early 1950s, Karczemne Lake was transformed into a receiver for domestic sewage, as well as sewage from dairies, slaughterhouses, breweries, furniture factories and municipal hospitals. Over 30 years, 60% of the raw municipal sewage went to Karczemne Lake through six sanitary sewers. With the expansion of the urban area, the amount of municipal sewage increased. Impurities from emerging single-family housing were collected in leaking cesspits, which caused some of the pollution to seep into the ground and migrate toward the lake. Some cesspits had overflows to the combined sewerage network, and municipal sewage flowed directly to the lake in this way. It was also possible to illegally connect sanitary sewers from individual domestic properties to the stormwater drainage system and directly discharge sewage to the lake from homes located along the shore. The improvement of water and sewage management in Kartuzy City began in the mid-1970s. In 1982, a mechanical-biological sewage treatment plant was opened. In the 1990s, almost the entire city was connected to a sanitary sewerage system, and it was only in 2010 that the management of stormwater began to be organized through the construction of settling tanks and stormwater separators. Until 2018, the stormwater drainage network in Kartuzy covered only part of the city—17%. This situation meant that during heavy rainfall, local flooding occurred, and the excess rainwater and snowmelt from streets that did not have a rainwater drainage system penetrated the sanitary sewer system, overloading it. The connection of the existing stormwater drainage system to the sanitary sewer system caused raw sewage to load into the lake.

The total catchment basin of Karczemne Lake covers 5.15 km2. This area is covered by two entirely different (in their ability to activate the load of nutrients in a surface flow) forms of land use: forests (57%) and urbanized areas (43%). The direct catchment of Karczemne Lake, excluding the area covered by the drainage system collecting rainwater, covers 0.45 km2: wasteland constitutes 20.7% of its total surface area, and forests grow over the other 79.3%. The calculations of the amount of nutrients that are annually brought to the lake with a watercourse or from outfalls were performed based on their actual concentration in water (total phosphorus—TP and total nitrogen—TN) and flows measured at the individual stations during yearly field studies. The partial load for a given day was the product of the volume of water (momentary flow) and the concentration of given nutrients in the water. Nutrient loads were calculated with a generally accepted method of time periods. The magnitude of a lake load with nutrients originating from surface flows from the direct catchment was calculated with a method that is recommended and applied by the OECD (Organisation for Economic Cooperation and Development) that involves the use of flow coefficients. The load of nutrients introduced to a lake with precipitation was determined based on the coefficients of pollution deposition per surface unit. It was assumed that angling baits are the main sources of pollution resulting from the recreational use of lakes. Based on the data obtained from Group No 57 of the Polish Angling Association from Kartuzy and conversion factors described by Wołos and Mioduszewska [26], the average number of anglers who used baits was calculated for each of the lakes. It was also assumed that the average content of biogenic compounds in fish feed was 3.0 g P and 12.0 g N per kg of bait.

Until the end of 2018, Karczemne Lake was a receiver of nutrient loads from both the catchment and atmosphere, which involved the following basic components: areal sources; inflow of waters via a watercourse that links Karczemne Lake with Mielenko Lake (this watercourse flows into the reservoir in the central part of the west bank); point sources (6 stormwater outfalls and sometimes the illegal discharge of municipal sewage); atmospheric sources and recreation, such as angling. The total loads of phosphorus and nitrogen introduced into Karczemne Lake were 134.7 kg P year−<sup>1</sup> (0.330 g P m−<sup>2</sup> year−1) and 1133.8 kg N year−<sup>1</sup> (2.80 g N m−<sup>2</sup> year−1), respectively. The permissible and dangerous (critical) loads calculated for this lake from Vollenweider's hydrological model [27] were 0.030 g P m−<sup>2</sup> year (12.1 kg year<sup>−</sup>1) and 0.060gPm−<sup>2</sup> year (24.2 kg year<sup>−</sup>1), respectively. The analysis revealed that the total phosphorus load introduced from external sources to the lake exceeded the dangerous load and was responsible for accelerated eutrophication. The phosphorus load introduced to Karczemne Lake was 550% higher than the critical load.

### *2.2. Water Sample Collection and Analysis*

The physicochemical properties of water samples from Karczemne Lake were determined over an annual cycle (April 2018, June 2018, August 2018 and November 2018). Samples were taken at the point of the maximum depth (Figure 2). The scope of the water analysis included the total phosphorus (TP; standard methods 2012), total nitrogen (TN; Shimadzu TOC 5000 analyzer, Kyoto, Japan), chlorophyll a by the colorimetric method (after concentration on a glass fiber filter Whatman GF/B and extraction with acetone—Nanocolor UV/vis, Macherey-Nagel (GmbH & Co., KG, Frankfurt, Germany); 750/664 nm before and 750/665 nm after acidification) and visibility (Secchi disc).

**Figure 2.** The spatial distribution of research points with the thickness of the sediment layer meant to mining.

The trophic state index (TSI) was calculated based on the concentrations of total phosphorus, total nitrogen and chlorophyll a, as well as Secchi disc visibility [28,29].

The coefficient of variation (CV) for the repeated analysis was 2% [30].

### *2.3. Bottom Sediment Collection*

The bottom sediment sampling was carried out on 7, 8, 9 and 11 March 2018. Samples for analysis were taken at 28 research stations (Figure 2), for which geographical coordinates were determined (Table 2).

The sediment cores were obtained using a Kajak tube sampler (KC Company, Silkeborg, Denmark). Every core length included the sediment layer down to the parent rock (floor). In this way, the thickness of the retrieved sediments was determined precisely at individual sites, and their organoleptic properties were assessed. During the collection, individual cores were divided into layers with a thickness of 30 cm (a—0–30 cm, b—31–60 cm, c—61–90 cm, d—91–120 cm, e—121–150 cm and f—151–180 cm) and prepared as separate samples. A total of 112 material samples were collected from the bottom of the lake.


**Table 2.** Geographical coordinates of sediment sampling points from Karczemne Lake.

### *2.4. Bottom Sediment and Interstitial Water Sample Analysis*

The output, after being transported to the laboratory, was weighed to determine the volumes and densities of the fresh samples. The samples were then dried to constant weight, and the volume and density were redetermined. The interstitial water was obtained from the sediment by centrifugation at 3000 rpm for 20 min according to Brzozowska and Gawro ´nska [31].

In the interstitial water samples, ammonia was examined by ionic chromatography using an ICS-5000 DIONEX DC, TKN by IL 550 TOC-TN analyzer, Hach Inc. (Larimer Country, CO, USA), phosphates and total P were measured by a Nano color spectrophotometer, and organic P was calculated as the difference between the TP and phosphate amounts. In the investigated sediments, organic matter was measured by LOI (the percent weight loss during the ignition) using weight analysis [31], total nitrogen was measured by the Kjeldahl method and silica was measured using weight analysis after preliminary digestion of the sediment sample in a mixture of strong mineral acids (H2SO4, HClO4 and HNO3 1:2:3) and filtering through a No. 390 filter. The phosphorus fractions were analyzed by sequential extraction according to the Rydin and Welch method [32] modified by van Hullebush et al. [33]. In the analyzed sediments, samples were examined for the following:


Extracted phosphorus was determined using the molybdate blue method according to Hermanowicz et al. [34]. The results were statistically analyzed (basic statistics—mean values, standard deviation) using a Statistica 13.0 software package [35].

### *2.5. Studies of the Spatial Variability of Phosphorus and Nitrogen Concentrations in the Bottom Sediment*

Lake restoration using the dredge method must be planned very precisely. The most important element of the project is to precisely define the thickness of the deposits that should be removed from the ecosystem. The thickness needed to remove the deposits is determined on the basis of the changes in phosphorus content in the sediment and the fractions in which phosphorus is present. The analysis of the Karczemne Lake bottom was planned in such a way to obtain the most complete information about the structure of its bottom. Due to the lack of data regarding the thickness and properties of the bottom sediments, probes were planned at points evenly distributed over the whole bottom surface of the lake. This methodical approach is recommended by the EPA [36] in situations with limited preliminary data. The results of the soundings referred to the partial surfaces of the lake bottom for which the given probing point was the geometric center (Voronoi diagrams; Wolfram Math Worlds [37]).

### *2.6. Studies of Heavy Metals and Persistent Organic Pollutant Contents*

In accordance with the guidelines included in the Ordinance of Ministry of the Environment (OME) [38] on waste recovery outside installations and equipment, six aggregate samples were prepared for analyzing the content of stable organic pollutants (POPs), polycyclic aromatic hydrocarbons (PAHs) and heavy metals in the sediments.


Limitations on the further use of bottom sediments or other types for development can result from higher than normal heavy metal content according to Hermanowicz et al. [34] or the presence of extremely hazardous persistent organic pollutants (POPs), such as hexachlorobenzene (HCB), polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons (PAHs). Analyses of heavy metals, polycyclic aromatic hydrocarbons and polychlorinated biphenyls were performed by the AAS (Atomic Absorption Spectrometer), ICP-AES (Inductively Coupled Plasma-Atomic Emission Spectrometer) and GC MS (Gas Chromatography Mass Spectrometry) methods. The coefficient of variation (CV) for the repeated analysis was 2% [30].

The usefulness of deposits for development was verified based on The Act on Waste [39] and concentrations of toxic substances caused by pollution [39]. The waste catalog classifies bottom sediments as waste with code 17 05 05 (dredging spoil containing or contaminated with dangerous substances), with the note that the sediments are hazardous waste or 17 05 06 (dredging spoil other than those mentioned in 17 05 05).

### **3. Results**

### *3.1. Trophic Status*

The value of every index calculated from the concentrations of total phosphorus, total nitrogen, visibility and chlorophyll a exceeded TSI 70 (Figure 3). This demonstrated strong pollution of the analyzed lake.

**Figure 3.** The trophic status index (TSI) calculated on the basis of visibility (SD) chlorophyll concentrations (Chl), total phosphorus (TP) and total nitrogen (TN).

### *3.2. Phosphorus and Nitrogen Content in the Interstitial Water of Karczemne Lake*

The interstitial water of Karczemne Lake was rich in biogenic compounds. The mean content of total phosphorus varied between 0.61 and 10.0 mg P L−<sup>1</sup> (Figure 4).

**Figure 4.** Mean content of total phosphorus in the interstitial water of Karczemne Lake (mg P L<sup>−</sup>1).

In the composition of total phosphorus in the interstitial water of the deeper sediment layers, organic phosphorus predominated, while in the layer with a thickness of 30 cm, phosphates prevailed. The highest concentration of phosphorus was recorded in the interstitial water separated from the bottom sediment that was under the influence of anthropopression, i.e., near the outlet of stormwater

collectors that previously experienced the inflow of domestic and industrial sewage. Sewage discharge points were located along the eastern shore of the lake. There, the maximum concentration of total phosphorus usually occurred in the interstitial water of the b layer: 31–60 cm.

The mean total nitrogen content in the interstitial water of Karczemne Lake varied from 9.60 to 71.40 mg N L−<sup>1</sup> (Figure 5). The concentrations of TN in the interstitial water clearly increased with depth in the sediment. In the nitrogen structure, the organic form was predominant. In the analyzed water, mineral nitrogen was in the form of ammonium nitrogen, the amount of which varied between 4.00 (St. 2 a) and 37.80 mg N L−<sup>1</sup> (St. 27 e). The values of ammonium nitrogen increased in the interstitial water in the sediment. The highest ammonia contents were found in the interstitial water of the bottom sediment located in the area bounded by 2 m and 3 m isobaths and at probing point No. 5.

**Figure 5.** Mean content of total nitrogen in the interstitial water of Karczemne Lake (mg N L<sup>−</sup>1).

### *3.3. Density and Hydration of the Analyzed Bottom Sediment*

Spatially, the bottom sediments of Karczemne Lake were characterized by a heterogeneous structure and variations in thickness and degree of consolidation.

The maximum thickness of modern sediments was 1.8 m. The thickness of the deposit is related to the long-term supply of sewage, which results in clear shallowing of the lake bed. Littoral deposits were characterized by the smallest thicknesses, with an average of 60 cm.

The deposits with the largest thickness occurred in the region of the maximum depth of the lake and in the places of inflow of municipal and industrial sewage.

The bottom sediments of Karczemne Lake had a loose structure that was susceptible to resuspension (agitation). The sediment mean hydration in the 30 cm thick layer a fluctuated from approximately 94–95% (Figure 6).

More dense and compact structures were observed only in the lower layers of the cores, especially in the littoral zone (Figure 7). The boundaries of structures within the cores were not clearly marked. The floor layer, constituting the native soil, was gyttja, i.e., organic and mineral sediment comprised mainly of the remains of allochthonous organisms with an admixture of sands. The mean density of the deposits of Karczemne Lake ranged from 0.99 to 1.45 g cm<sup>−</sup>3, and on average, it was 1.05 g cm−<sup>3</sup> (Figure 7).

**Figure 6.** Mean hydration of the bottom sediment of Karczemne Lake (%).

**Figure 7.** Mean density of the bottom sediment of Karczemne Lake (g cm<sup>−</sup>3).

### *3.4. Chemical Composition of Bottom Sediment of Karczemne Lake*

The average chemical composition of the bottom sediments of Karczemne Lake was quite variable. The sediments that accumulated in the central part of the basin, up to a depth of 60 cm, can be considered mixed according to the Stangenberg nomenclature [40] because none of the components was present in amounts exceeding 50% of the dry mass. The organic matter in these sediments ranged from 15.8% to 43.5% of the dry mass. In the coastal regions of the lake and in deeper levels of sediment from the central part of the reservoir, the dominant component was silica, accounting for up to 92% of the dry mass. These bottom deposits can be described as silicate. The amounts of nitrogen in the analyzed sediments varied from 0.20% to 2.50% N dry weight (Figure 8a–d). Lower concentrations of nitrogen were found in the northern part of the lake along the axis determined by the inflow of the Klasztorna Struga (sounding stations 1–8). In the rest of the reservoir, the bottom deposits were richer in nitrogen, especially in the areas of the maximum depth of the basin.

The main components of the sorption complex of the bottom sediments of Karczemne Lake were iron (approximately 3.5% d.w.) and aluminum (approximately 3.0% d.w.). The combination of phosphorus and iron is unstable due to the variable valence of iron. Under anaerobic conditions, iron turns into a reduced form, releasing phosphates into the water. The bottom sediment of Karczemne Lake contained very high amounts of phosphorus, ranging from 0.1% to 7.3% of the dry mass (Figure 8a–d). Maximum amounts were found in the layer of deposits with a thickness of 60 cm and, at some points, up to 90 cm. Below 90 cm, much lower values were observed (Figure 8a–d).

The mean concentrations of total phosphorus in the bottom sediments of Karczemne Lake were very high. To date, no other lake in Poland has recorded such high values of phosphorus. The maximum amount of total phosphorus was found at research point number 5: 31.856 mg P g−<sup>1</sup> d.w. in the surface sediments of layer a (0–30 cm). Phosphorus amounts exceeding 10 mg P g−<sup>1</sup> d.w. were found at the following research points: 1 a (14.1 mg P g−<sup>1</sup> d.w.), 10 a (11.2 mg P g−<sup>1</sup> d.w.), 14 a (10.4 mg P g−<sup>1</sup> d.w.), 14 b (14.4 mg P g−<sup>1</sup> d.w.), 15 a (22.9 mg P g−<sup>1</sup> d.w.), 16 a (16.9 mg P g−<sup>1</sup> d.w.), 24 b (12.9 mg P g−<sup>1</sup> d.w.), 26 a, b (12.0 mg P g−<sup>1</sup> d.w.), 28 a (13.6 mg P g−<sup>1</sup> d.w.) and 28 b (16.6 mg P g−<sup>1</sup> d.w.; Figure 8a–d).

**Figure 8.** *Cont*.

**Figure 8.** (**a**) The mean chemical composition of bottom sediment of Karczemne Lake—Stations 1–7; (**b**) the mean chemical composition of bottom sediment of Karczemne Lake—Stations 8–14; (**c**) the mean chemical composition of bottom sediment of Karczemne Lake—Stations 15–21 and (**d**) the mean chemical composition of bottom sediment of Karczemne Lake—Stations 22–28.

### *3.5. Phosphorus Fractions in the Bottom Sediments*

The analysis of particular phosphorus percentages of the total phosphorus showed that the dominant types were NaOH-nrP (phosphorus associated with organic matter), which made up over 82% of TP at research point number 5, and calcium-related phosphorus (HCl-P), whose presence at points 24, 26 and 28 exceeded 45% of the TP. Residual phosphorus (res-P) made up 25% of the TP at some research points, and at point 14, it exceeded 35% in the 30–60 cm layer.

Additionally, the NaOH-rP fraction (phosphorus bound mainly with aluminum) at some sites occurred in higher amounts, exceeding 20% of the TP. These fractions are moderately biologically

available (NaOH-rP and NaOH-nrP), poorly available (HCl-P) or biologically inaccessible (res-P; Figure 9).

**Figure 9.** Share of phosphorus fractions (as %TP) at selected bottom sediment research stations.

The most mobile fractions of phosphorus—NH4Cl-P (phosphorus loosely bound to sediment) and BD-P (phosphorus sensitive to redox potential changes)—made up a small percentage of the total phosphorus (typically only a few percent). However, due to the extremely high content of total phosphorus in the surface layer of sediments, a few percent of the TP translates into very high mobile phosphorus content, especially in the case of the BD-P fraction; at several research points, the content of this fraction was higher than 2.5 mg P g−<sup>1</sup> d.w. (Figure 9). These values indicate a serious threat to the quality of Karczemne Lake water. In addition, NaOH-rP and NaOH-nrP were present in very high proportions in the studied sediments. These fractions are mobilized by the increased reaction in the interphase water. In a polluted reservoir such as Karczemne Lake, with its low depth and polymictic water dynamics, intense photosynthesis may affect the increase in pH in the sediment waters (interphase) and thus stimulate an internal supply of phosphorus from the NaOH-rP and NaOH-nrP fractions (Figure 9).

### *3.6. Stable Organic Pollutants (POPs), Polycyclic Aromatic Hydrocarbons (PAHs) and Heavy Metal Contents in the Sediments*

In sample No. 1, taken from research points 1–15 from layers a (0–30 cm) and b (31–60 cm), exceedances were found in the content of benzo (a) pyrene—1.21 mg kg−<sup>1</sup> d.w. in the ratio to the standard of 1.0 mg kg−<sup>1</sup> d.w.; benzo (b) fluoranthene—1.88 mg kg−<sup>1</sup> d.w., compared to the norm of 1.5 mg kg−<sup>1</sup> d.w.; benzo (g, h, i) perylene—1.44 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w. and indeno (1, 2, 3-cd) pyrene—1.75 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w. (Table 3).

In sample No. 3, taken from research points 16–28 from layers a (0–30 cm) and b (31–60 cm), exceedances were found in the content of benzo (a) pyrene—2.10 mg kg−<sup>1</sup> d.w. in the ratio to the standard of 1.0 mg kg−<sup>1</sup> d.w.; benzo (a) anthracene—1.69 mg kg−<sup>1</sup> d.w., compared to the norm of 1.5 mg kg−<sup>1</sup> d.w.; benzo (b) fluoranthene—3.57 mg kg−<sup>1</sup> d.w. in ratio to the standard—1.5 mg kg−<sup>1</sup>

d.w.; benzo (g, h, i) perylene—2.65 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w. and indeno (1,2,3-cd) pyrene—1.86 mg kg−<sup>1</sup> d.w., compared to the standard of 1.0 mg kg−<sup>1</sup> d.w.


**Table 3.** Organic pollutants and heavy metal contents in bottom sediment of Karczemne Lake.

In sample No. 5, taken from research point 5 from layers a (0–30 cm) and b (31–60 cm), exceedances were found in the range of benzo (a) pyrene—9.75 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w.; benzo (a) anthracene—13.0 mg kg−<sup>1</sup> d.w., compared to the norm of 1.5 mg kg−<sup>1</sup> d.w.; benzo (b) fluoranthene—14.4 mg kg−<sup>1</sup> d.w., compared to the norm of 1.5 mg kg−<sup>1</sup> d.w.; benzo (g, h, i) perylene—7.34 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w.; benzo (k) fluoranthene—5.77 mg kg−<sup>1</sup> d.w., compared to the norm of 1.5 mg kg−<sup>1</sup> d.w dibenzo (a, h) anthracene—3.24 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w.; indeno (1, 2, 3-cd) pyrene—8.03 mg kg−<sup>1</sup> d.w., compared to the norm of 1.0 mg kg−<sup>1</sup> d.w.; Cu—247 mg kg−<sup>1</sup> d.w., compared to the standard of 150 mg kg−<sup>1</sup> d.w. and Pb—317 mg kg−<sup>1</sup> d.w., compared to the standard of 200 mg kg−<sup>1</sup> d.w.

In samples No. 2, 4 and 6, exceedances beyond the limit values were not found (Table 3).

With reference to the regulation, 40% of the bottom sediment of Karczemne Lake did not meet the requirements of uncontaminated spoil (code in the waste catalog 17 05 06) and constituted a dangerous sediment, i.e., catalog code 17 05 05, and should be disposed by incineration. Sixty percent of the output could be transferred for further use after prior dilution, i.e., mixing with sediments formed during the wastewater treatment process. In the case of introducing it to the ground (leveling, reclamation and fertilization), the provisions of the OME [38] on the method of conducting the assessment of the ground surface pollution indicate that dredged spoil on the surface of the earth cannot make the original classification of the soil or soil quality change as a result of its deposit.

### *3.7. Mass Calculation*

The results of the research on the spatial variability of phosphorus content and its fractions in the Karczemne Lake sediments enabled the precise determination of sediment layers that need to be removed from particular sectors of the ecosystem bottom. The deposit thicknesses indicated for removal in each of the separated bottom sediment sectors are presented in Table 4. Considering the sector areas and the thickness of the layer to be removed, the volume of deposits to be dredged from Karczemne Lake was 240 013 m3 (Table 4). Based on the density of the sediment and the degree of hydration, it was estimated that the fresh mass of the output will be 2,588,054.7 Mg and that the dry weight of the output will be 2691.35 Mg (Table 4). The amount of nutrients that will be removed from the ecosystem along with the sediments and interstitial waters will be 212.86 Mg of phosphorus and 405.67 Mg of nitrogen (Table 5).


**Table 4.** Characteristics of physical parameters of bottom sediments with respect to the spatial volume balance and the mass of potential output to be removed.

**Table 5.** The load of phosphorus and nitrogen can be withdrawn from the Karczemne Lake with bottom sediments and interstitial water.


### **4. Discussion**

The results of monitoring have shown that the water and bottom sediment of Karczemne Lake should be classified at the "below good" level. The lake does not meet the requirements of the Water Framework Directive of the European Union [41]. With such high nutrient concentrations and excess organic matter, the natural inhibition of the severe degradation of the lake is impossible. In addition to the very poor water quality, the bottom sediments of the analyzed lake also contain extremely high concentrations of pollutants. The concentrations of total phosphorus in the bottom sediments of Karczemne Lake were very high compared to those of other lakes [42–44]. To date, no other lake in Poland has recorded such high values of phosphorus. The maximum amount of total phosphorus was found at research point number 5: 31.8 mg P g−<sup>1</sup> d.w. in the surface sediments of layer a (0–30 cm). Bojakowska [45] reported that in Karczemne Lake sediments, a maximum of 1.9% TP in dry matter was found, which corresponds to 19 mg P g−<sup>1</sup> d.w., while previous studies carried out by Grochowska et al. [46] showed a maximum of just over 12 mg P g−<sup>1</sup> d.w. in the surface layer of sediments (10 cm thick) taken from the deepest research point. For comparison, the maximum phosphorus content found by Kentzer [47] in the bottom sediment of dystrophic Zmarłe Lake was slightly above 10 mg P g−<sup>1</sup> d.w. In the sediments of Długie Lake in Olsztyn, which was polluted for many years by domestic sewage, the level of phosphorus in the bottom sediments ranged from 4 to 6 mg P g−<sup>1</sup> d.w. [48]. According to Sahin et al. [49], the amount of phosphorus in the sediments ranged from approximately 0.5 to 20.2 mg P g−<sup>1</sup> d.w. However, Augustyniak [50] suggested that in the sediments of shallow lakes were not contaminated by sewage, the phosphorus content is usually low and does not exceed 2–3 mg P g−<sup>1</sup> d.w. Karczemne Lake was a domestic sewage receiver for over 60 years. The pollution loads introduced with the sewage inflow deposited on the bottom of the reservoir forming a layer of so-called "modern sediments". In oligotrophic lakes, the thickness of the sediment increases by 1 mm each year, and in hypertrophic lakes, it increases by approximately 1 cm [50]. Such high phosphorus content in the bottom sediments, especially in the surface layer, is certainly the result of massive sewage pollution taking place over the last several decades. Station 5, with the highest level of phosphorus, is located in the immediate vicinity of the collector discharging sewage into the lake, which for many years was the main source of pollution. It has been reported that in the deepest layers of the sediment (from 60 cm and in some positions from 90 to 150 cm deep into the sediment), the recorded amounts of total phosphorus are from approximately 0.5 (St. 5) to approximately 3 mg P g−<sup>1</sup> d.w. (St. 18), which gives an overview of the concentration of total phosphorus before the pollution of the lake. Due to the low depth and polymictic nature of the reservoir bottom, sediment certainly undergoes resuspension, which in some sites could cause higher phosphorus contents in the deeper layer of sediment (30–60 cm), e.g., stations 24, 26 and 28. The vertical profiles of the phosphorus and total phosphorus fractions in the sediments at selected research stations show very rich concentrations of phosphorus in more detail as the deposit depth increases in Karczemne Lake. Theoretically, this phosphorus can easily be mobilized, providing a serious load of pollutants to the water of the analyzed reservoir.

### *4.1. Reduction in Pollution Loads by Protective Methods*

Currently, the sewage network of the city of Kartuzy is being modified, which is necessary to decrease the high nutrient loads from wastewater. In addition, renovation work is being carried out in the tributary to Karczemne Lake (Klasztorna Struga) and involves the regulation of the riverbed and the fascine. In-stream and riparian vegetation will be restored in the channel. The aboveground parts of hydrophilic plants assimilate biogenic elements and increase aesthetic value [51]. In turn, the underground parts of plants (rhizomes and roots) release oxygen into the rhizosphere, which supports the processes of biodegradation of organic matter and nitrification, as well as biogenic substances. *Glyceria maxima*, *Phalaris arundinacea* and *Phragmites australis* will be planted in the channel of the Klasztorna Struga. The last protective action is the prohibition of angling baits.

Applying all of the aforementioned solutions will result in a reduction in the external load of the lake by approximately 80%.

### *4.2. Conception of Restoration Treatments*

In the past, Karczemne Lake was overloaded by pollution from the catchment, which was a result of the long-term input of raw wastewater, including domestic, sanitary, storm and industrial waste. Under these conditions, it is necessary to use the optimal method of lake restoration after protection techniques are implemented in the catchment. The optimal restoration method for Karczemne Lake will be the removal of deposits due to their unusually high concentration of phosphorus. Additionally, for the stabilization of environmental conditions in the reservoir, the phosphorus inactivation method and biomanipulation will be applied.

For the purpose of restoration treatments in Karczemne Lake, there are separate extraction fields (10) in which the outputs of the liquid and solid fractions will be successively carried out for removal. The field currently undergoing the deposit removal process will be separated from the rest of the lake by a plastic curtain to prevent the movement of the deposit. The removal of the deposits will be carried out by a method with the use of an innovative device, the only type of equipment in Europe (suction dredge with wormwheel), precisely designed for lake dredging. Extracted output will be pumped to a pipeline on the northern shore of the lake and later by hydrotransport to a sewage treatment plant.

This technical solution guarantees prevention of the uncontrolled runoff of leachate back into the lake basin. The pipeline will be made of plastic (Figure 10).

**Figure 10.** The scheme of planned restoration.

The proposed method does not require the construction of a completely new deposit treatment system consisting of presses, centrifuges, polymer-dosing stations, water-conditioning equipment and reaction pools (Figure 10).

In the sewage treatment plant, the sediment will be dehydrated in centrifuges and then subjected to hygienization. The part of the sediment that is contaminated with heavy metals will be transported for utilization.

Unpolluted sediments extracted from the lake and processed at the sewage treatment plant can also be valuable material for improving the soil structure, as they may contain premineralized organic compounds and easily digestible fertilizers, including macro- and microelements such as nitrogen, 1.5% N; phosphorus 3, 5% P2O5; calcium, 4.22% CaO; magnesium, 1.6% MgO and iron, 1.8% Fe2O3

(average contents of elements in dry mass of the Karczemne Lake sediments). Another form of deposit utilization can be the reclamation of degraded areas, for example, landfills [52–54].

In relation to the current level of external loads, the quantities of phosphorus and nitrogen that will be extracted from the lake during reclamation (loads coming with surface inflow, outlets of rainwater collectors, recreation and atmospheric precipitation) correspond to the amount of pollution that reaches the lake in 1300 years for phosphorus and 300 years for nitrogen. This confirms the validity of the assumptions behind the reclamation methods adopted, for which the priority was to remove the bottom sediments outside the ecosystem.

The removal of the bottom sediments from Karczemne Lake will be carried out within two years, from spring to the moment of the lake ice cover.

In the extraction field in which the sediment will be removed, the first phase of phosphorus inactivation to chemically precipitate pollutants distributed in the water column as a result of the dredging will be carried out. After the entire lake basin has been dredged, the next four stages of phosphorus inactivation (spring, autumn, spring and autumn) will be carried out. Based on the experience and research results obtained by the Department of Water Protection Engineering in the field of lake reclamation by phosphorus inactivation [2,7,55,56], two types of coagulants will be dosed: PAX 18 and PIX 111. The doses of PIX and PAX coagulants were determined on the basis of the phosphorus fractions in the water and bottom sediments and the amount of natural components in the sediments having sorption capacity in relation to phosphorus, such as Al, Fe and Ca [57,58].

The calculated doses of aluminum coagulant for Karczemne Lake were as follows: the area of the central zone of the lake was 20.0 ha, the volume of the central zone was 660,000 m3, the demand of the profundal sediment for aluminum was 23.6 g m<sup>−</sup>2, the demand for aluminum with which to bind the phosphorus in the water column in the central zone was 1650 kg and the total amount of reactive aluminum was 6370.8 kg. The total dose of PAX 18 was 70,787.0 kg.

Next, the dose of the iron coagulant for Karczemne Lake was calculated as follows: the area of the coastal zone was 20.4 ha, the volume of the coastal zone was 378,000 m3, the demand of the littoral deposits for iron was 44.26 g m<sup>−</sup>2, the demand for iron for binding the phosphorus in the water depth in the coastal zone was 945 kg and the total amount of reactive iron was 9974.0 kg. The total dose of PIX 111 was 74,433.0 kg.

To maintain the positive effects of restoration in Karczemne Lake, rational fish farming should be utilized. This method can be implemented through the use of biomanipulation, which involves the conscious shaping of the biocenosis of aquatic organisms by changing the species composition of ichthyofauna. The main goal of biomanipulation is to increase the number of large forms of zooplankton, mainly Cladocera, and through their control, feed the amount of phytoplankton, which are the first link in the trophic pyramid, thus reducing water blooms.

To achieve this effect, the number of sedentary fish, such as bleak, sunflower, small perch, juvenile roach, silverfish, bream and crucian, which feed on zooplankton or seek food in bottom sediments, should be limited. The reduction in these fish assemblies can be achieved, inter alia, through selective harvesting (without the use of towed tools). According to Gołdyn [59], at least 75% of the initial fish should be harvested. This treatment should be intense and short term and should optimally occur within one year but no later than two years. The best effects result from the use of a large amount of pike fry. The recommended amount of recessed fry is at least 1000 pcs/ha [59], and the restocking material should have a size exceeding 10 cm because at this stage of development, the pike is going to feed on fish.

### **5. Conclusions**

After detailed monitoring of the water and bottom sediment of Karczemne Lake, the methods proposed for protection and restoration of the lake discussed above enable us to realistically assess the potential opportunities for the renewal of severely degraded shallow, urban lakes. Considering the chemical composition of the water and bottom sediments of Karczemne Lake, external conditions, long-term economic and environmental aspects and the development of new technologies, it appears that an optimal variant of the approach taken to achieve permanent water quality improvement might consist of modernization of the urban sewerage system and restorative actions, including the removal of the bottom sediments from Karczemne Lake by a completely new, safe and economically justified method of mining and managing bottom sediments and phosphorus inactivation along with supportive actions in the form of biomanipulation. The technology of bottom sediment removal along with the construction of a hydrotransport pipeline, the phosphorus inactivation method and biomanipulation will have a total cost of US \$2,100,000.

**Author Contributions:** Conceptualization J.G. and M.Ł., investigation J.G., M.Ł., R.T., R.A., verification R.A. and R.T., writing-original draft preparation J.G. and R.A., writing-reviewing and editing J.G., R.A., M.Ł., R.T. All authors have read and agreed to the published version of the manuscript.

**Funding:** "Project financially supported by Minister of Science and Higher Educationin the range of the program entitled "Regional Iniciative of Excellence" for the years 2019—2022, project No. 010/RID/2018/19, amount funding 12.000.000 PLN".

**Acknowledgments:** Authors thank the Community of Kartuzy.

**Conflicts of Interest:** Authors declare no conflict of interest.

### **References**


© 2020 by the authors. Licensee MDPI, Basel, Switzerland. This article is an open access article distributed under the terms and conditions of the Creative Commons Attribution (CC BY) license (http://creativecommons.org/licenses/by/4.0/).

### *Review* **Developments in the Use of Lipase Transesterification for Biodiesel Production from Animal Fat Waste**

**Fidel Toldrá-Reig 1,**†**, Leticia Mora <sup>2</sup> and Fidel Toldrá 2,\***


Received: 30 June 2020; Accepted: 21 July 2020; Published: 23 July 2020

**Abstract:** Biodiesel constitutes an attractive source of energy because it is renewable, biodegradable, and non-polluting. Up to 20% biodiesel can be blended with fossil diesel and is being produced and used in many countries. Animal fat waste represents nearly 6% of total feedstock used to produce biodiesel through alkaline catalysis transesterification after its pretreatment. Lipase transesterification has some advantages such as the need of mild conditions, absence of pretreatment, no soap formation, simple downstream purification process and generation of high quality biodiesel. A few companies are using liquid lipase formulations and, in some cases, immobilized lipases for industrial biodiesel production, but the efficiency of the process can be further improved. Recent developments on immobilization support materials such as nanoparticles and magnetic nanomaterials have demonstrated high efficiency and potential for industrial applications. This manuscript reviews the latest advances on lipase transesterification and key operational variables for an efficient biodiesel production from animal fat waste.

**Keywords:** biodiesel; fuel; energy generation; lipase; immobilized lipase; animal waste; lard; tallow; animal fat; transesterification

### **1. Introduction**

Animal byproducts generated in the European Union slaughterhouses represent nearly 17 million tons per year and, from them, 5 million tons inedible byproducts result from rendering and are mostly used for energy generation like biofuels and biodiesel [1–3]. After rendering byproducts, fat is obtained from beef tallow, mutton tallow, pork lard and chicken fat [4,5]. Such fat is majorly composed of triacylglycerols with fatty acids of 16 to 18 carbons. The most abundant saturated fatty acids are palmitic (16:0) and stearic (18:0) acids; the major monounsaturated fatty acid is oleic acid (18:1) and the most abundant polyunsaturated fatty acids are linoleic (18:2) and arachidonic (20:4) acids [6,7]. Animal fat waste is also obtained from the meat processing industry and from recycled waste from the cooking business [8,9] that are classified as yellow grease if the content of free fatty acids is lower than 15% by weight and brown grease when it is higher than 15% [10]. In 2019, more than 800 thousand tons of animal fats, equivalent to 6% of total feedstock, were used to produce biodiesel in the European Union [11,12], while 8.4% of total feedstock was used in the case of the US, consisting of mainly 74 tons of poultry fat, 132 tons of tallow and 243 thousand tons of white grease [13].

Biodiesel produced from animal fats is cheaper than when made from vegetable oils. An additional advantage is that fossil CO2 reduction is higher when using animal fat for biodiesel generation; nearly 80% CO2 reduction may be reached for animal fat in comparison to 30% reduction when using vegetable oil [14,15]. The bioenergy demand is continuously increasing and in 2050 it is expected

to reach 30% of the fuel consumed in the world for road transport [15,16]. Research on biodiesel production is trying to maximize the yield and minimize the costs by using better catalysts that can be reused and improve the transesterification efficiency [17,18]. Furthermore, the feedstock used as raw material for biodiesel production represents up to 80% of the total cost [19] and it explains its variability in different geographic areas depending on the climate and agriculture [20].

Total biodiesel world production has been increasing progressively year by year, reaching nearly 45 million tons in 2019 [12]. The European Union has the largest biodiesel production through its 202 plants producing more than 14 million tons of biodiesel in 2019 [11,21]. More than 5.6 million tons of biodiesel were produced in the US in 2019 through its 91 plants [13,22]. Nearly 80% of new diesel vehicles are prepared for B20 use that consists of fossil diesel blended with 20% biodiesel [13].

Transesterification through alkaline catalysis is the preferred process at industrial biodiesel production plants [23]. However, raw materials like animal fat that contain moisture and free fatty acids are troublesome for alkaline transesterification due to soap formation. Acid catalysis does not have such troubles, but the reaction is much slower than alkaline catalysis, needs a larger size reactor and requires a higher alcohol to fat molar ratio [24]. Heterogeneous catalysts are not sensitive to the presence of free fatty acids and moisture, can catalyze esterification and transesterification simultaneously, and can be separated from the reaction media. However, such solid catalysts tend to form three phases resulting in a reduced reaction rate and high energy consumption [25]. The simultaneous esterification and transesterification also occur with supercritical technology where high temperature and pressure conditions (i.e., >250 ◦C and 10 MPa) increase the solubility and reduce the mass transfer limitation resulting in good efficiency but with high energy consumption [26–28]. Pseudo catalytic transesterification using biochar as the porous material for the pseudo-catalytic reaction at more than 300 ◦C has the same advantages as supercritical transesterification, but also has high energy consumption [29,30]. Therefore, animal fats may be processed for biodiesel production through enzymatically catalyzed transesterification even though some issues, like the cost of lipase and its poor stability, can be improved through immobilization [31]. Lipases have the advantage to generate biodiesel under mild reaction conditions through the conversion of free fatty acids and triacylglycerols in the presence of an acyl acceptor [32]. This manuscript reviews and discusses the latest advances in the use of free and immobilized lipases for an efficient transesterification of animal fat waste.

### **2. Mechanisms of Action of Lipases**

Lipases, triacylglycerol ester hydrolases (EC 3.1.1.3), are serine hydrolases with an active site containing an amino-acid triad of serine, histidine and aspartate [32]. Lipases are obtained from a variety of sources such as animal and plant tissues and microorganisms. Lipases show a wide range of pH and temperature for activity and vary from strain to strain regarding specificity and hydrolysis rate [33] Lipases exhibit good stability in non-aqueous mediums and exhibit neutral pH range; such stability is increased when the enzyme is immobilized.

Lipases can catalyze esterification, inter-esterification, and trans-esterification reactions in non-aqueous environments. Lipases catalyze the hydrolysis of triacylglycerols at the aqueous-non aqueous interface but these enzymes can also catalyze the synthesis of esters from alcohols and long chain fatty acids in low moisture environment [33]. Lipases follow a two-step mechanism for the generation of fatty acid methyl esters in transesterification reactions, usually through the Ping-Pong Bi Bi mechanism [34].

Most triacylglycerol lipases are regiospecific because they can only hydrolyze primary ester bonds at the sn-1 and sn-3 positions, external positions within the triacylglycerol, and can generate either one free fatty acid and diacylglycerol, or two free fatty acids and 2-monoacylglycerol that remain unhydrolyzed. The full process from triacylglycerols into biodiesel and glycerol as end products is shown in Figure 1. Regiospecificity is characteristic of extracellular bacterial lipases from *Bacillus* sp. [35,36].

Monoacylglycerol lipases (EC 3.1.1.23) catalyze the hydrolysis at the specific sn-2 position of 2-monoacylglycerol into free fatty acid and glycerol. Such lipases may be present in the enzyme extract and masked when measuring activity with standard activity methods like those based on triolein hydrolysis measurement. Monoacylglycerol lipases have been the object of few studies [37], although they might be present in some microbial enzyme preparations [38]. Other lipases are nonspecific and can act on any of the ester bonds of the triacylglycerol and therefore break down the triacylglycerol to release free fatty acids and glycerol as the final products. This is the case of lipases from *Staphylococcus aureus* and *hyicus* [39], *Geotrichum candidum*, *Corynebacterium acnes*, *Penicillium cyclopium* [24] and *Chromobacterium viscosu* [40]. Another alternative for the hydrolysis of monoacylglycerols is the acyl migration in the glycerol backbone from the sn-2 position to sn-1 or sn-3 positions [41].

The specificity of lipases depends on the length of fatty acids, presence of double bonds, branched groups and, consequently, reaction rates may have important variations depending on the composition of triacylglycerols present in the fat waste. Lipases are especially active against medium to long chain fatty acids, which are those more usual in animal fat waste [17].

**Figure 1.** Transesterification of animal fat to biodiesel. TGL: triacylglycerol lipase; nsTGL: non specific triacylglycerol lipase; MGL: monoacylglycerol lipase.

### **3. Sources of Lipases**

Most lipases originated from microorganisms and are produced in fermenters under controlled conditions (see Table 1). Lipases are produced by a variety of gram-positive and gram-negative bacterial strains, especially from the genera of *Pseudomonas* [42,43], also by filamentous fungus that are commercially important such as those belonging to the genera of *Rhizopus* sp. [44], *Aspergillus* sp. [45], *Penicillium* sp. [46], *Geotrichum* sp. [33], *Mucor* sp. [47] and *Thermomyces* sp. [48]. Lipases produced from yeasts are also relevant such as those from *Candida* sp. [49,50].

Extracellular lipases are secreted into the production medium and recovered from the microorganism broth. Then, lipases are further separated and purified but downstream processing is costly. Intracellular lipases imply the use of whole cell microorganisms and this fact reduces the costs of enzyme extraction and purification but the efficiency and biodiesel yield is low when catalyzing an oily substrate due to mass transfer limitations for substrate penetration and product release [28,51]. Some whole cell biocatalysts used to produce biodiesel are filamentous fungi like *Aspergillus* and *Rhizopus* [49].


**Table 1.** Bacteria, yeasts and filamentous fungi producing lipase and sources of isolation.

### **4. Free Lipase**

Lipases constitute an attractive catalyst for transesterification in those wastes containing large amounts of moisture and free fatty acids, which is the case of animal fat and is what makes it troublesome for alkaline transesterification. Table 2 shows some examples of the use of free lipases for biodiesel production from animal fat waste. The use of lipases has relevant advantages over conventional alkali catalysts. The most relevant are the absence of soap formation in the reactor, insensitivity to water content and acidity value, moderate reaction conditions, broad substrate range, good purity of biodiesel after transesterification and absence of pollutants, especially when treating cooking oil waste containing large amounts of free fatty acids [75]. On the contrary, there are also important disadvantages such as the high enzyme costs, poor enzyme stability, and the enzyme deactivation by alcohol [76] and partly by the generated glycerol [66].

As said, lipases are sensitive to the alcohol, in most cases methanol, used for biodiesel production and this fact increases the operational costs. There are some alternatives to avoid enzyme damage by methanol: stepwise addition of methanol to reaction mixtures avoiding a high concentration [65], the use of co-solvents like hydrophilic tert-butanol that dissolve glycerol and methanol and therefore allow high transesterification yields and rates [77], also the addition of longer-chain alcohols [78], or methyl or ethyl acetate as acyl acceptors [79,80]. Another solution is the use of novel lipases that can support one-step addition of high methanol concentration [81]. In this sense, another solution is the use of tools like protein engineering, recombinant methods and metabolic engineering that are used to improve thermostability as well as stability in organic solvents [80]. Glycerol may be extracted with organic solvents although the enzyme activity may be affected [81].

Lipase transesterification requires an extended time of reaction, and has a slow conversion rate as shown in Table 2. The recovery of the enzyme is rather difficult and the enzyme stability is poorer at high temperature and pH [81–84]. All these troubles have hindered its adoption at an industrial scale and therefore, transesterification with alkaline catalysis is still preferred at biodiesel-producing industrial plants [23]. However, such troubles experienced with enzymes can be partly overcome through its immobilization on a solid material that acts as enzyme carrier and increases its stability and efficiency [81,85,86]. Immobilization also allows an easy downstream separation from the product and decreases cost [82]. However, recyclability is an issue because lipases tend to lose activity after continuous operation [86]. In any case, it was reported that the use of soluble lipases might be more

competitive if the commercial enzymes would have a price 50 times lower than the immobilized lipase [25,87]


**Table 2.** Biodiesel production with various free soluble lipases.

<sup>a</sup> Biodiesel yield (wt/wt.%) was determined as the methyl esters amount produced by the lipases in the reaction process divided by the initial amount of esters; or <sup>b</sup> by the amount of oil.

### **5. Immobilized Lipase**

### *5.1. Types of Supports and Immobilization Procedures*

The immobilization of lipases consists of the retention of the enzyme at the surface of the support material. In this way, immobilized lipases show an improved efficiency and reduced costs, with longer enzyme stability and better resistance to denaturation by alcohol. There are many available supports of organic, synthetic and inorganic nature for lipase immobilization. Such materials may vary in characteristics such as particle dimensions, shape, pore volume, hydrophobicity, and density and must be stable to physical, chemical, and microbial degradation [93]. Porous supports with controlled pore distribution are very interesting for lipase immobilization because they offer an extensive surface area and therefore, higher enzyme loading. However, caution must be observed if pores are too small because they could get blocked by the enzyme, reducing its efficiency.

There is a large variety of immobilization procedures (see Figure 2) such as adsorption, covalent binding, cross-linking, entrapment, or encapsulation that have been developed to enhance the catalytic activity, and its stability, and make possible the reutilization of the enzyme in relation to the soluble lipase [94]. Some methods like cross-linking enzyme aggregates are not considered in this review because even though they are inexpensive, highly efficient, and do not need support for immobilization, they have rather poor mechanical stability [95].

Immobilization by adsorption on materials such as water-absorbing polymer, hydrophobic macroporous polypropylene particles or silica gel is simple, but can result in an undesirable leakage making it necessary to assure the retention of the enzyme by additional ionic or covalent bonds. This can be an ion exchange resin or cross-linking with glutaraldehyde [68]; performance was improved by crosslinking with glutaraldehyde. The stability in acid pH was improved as well as thermostability at 45–50 ◦C. It retained 80% of relative biodiesel production after 5 consecutive batches [96]. Adsorption of lipase from *Burkholderia cepacia* was compared to covalent immobilization on epoxy acrylic resin. The adsorbed enzyme gave a higher conversion than the covalent one after a three-step addition of ethanol, 68% vs 47% [97]. The covalently immobilized enzyme showed lower affinity towards diglycerides and monoglycerides; this was attributed to blockage of the active groups by the covalent bonds to the support material, which resulted in enzyme rigidity [98]. Mesoporous materials are attractive because they have high surface area, larger pore volume, absence of toxicity, and good stability [99]. Examples of reported immobilized lipases used in recent studies are shown in Table 3.

**Figure 2.** Major types of biological catalysis for biodiesel generation.


**Table 3.** Biodiesel production with various immobilized lipases.

<sup>a</sup> Biodiesel yield (wt/wt.%) was determined as the methyl esters amount produced by the lipases in the reaction process divided by the initial amount of esters, or <sup>b</sup> by the amount of oil.

There is a good affinity for immobilization of lipases on hydrophobic supports [104], giving a fast and good attachment by hydrophobic adsorption [105]. Transesterification of waste lard was tested with immobilized lipase B from *Candida antarctica* with the assistance of ultrasound for improving the dispersion and collision of the reagent molecules. Ultrasonic wave amplitude, ultrasonic cycles, and reaction parameters were optimized and a kinetic model was developed [69]. Pulsed ultrasound irradiation increased by about 3 times the synthesis rate of fatty acid ethyl ester by lipases immobilized on hydrophobic carriers like octadecyl-sepabeads [106].

Porous silica nanoflowers have center-radial pore structure that allows the load of lipase inside the structure to have good mass transfer for substrates and products [107]. Dichlorodimethylsilane was used to modify the silica nanoflowers for the adsorption of *Candida antarctica* lipase B and the biocatalytic pickering emulsion was constructed [94]. Pickering emulsion stabilized by hybrid nanoparticles [108], solid particles [109], or carbon nanotubes crosslinked with lipase [110], have been constructed and successfully used [111,112]. In this way, this emulsion facilitates biphasic reactions and simplifies the recovery of lipase that remains in its microenvironment [113].

Metal-organic frameworks provide advantages for immobilization: they can be easily separated, they offer an extended surface area that can be tuned, they have adequate pore size, have structural and functional diversity and good stability. Immobilization may be through adsorption, encapsulation, and coprecipitation [114]. Lipase is strongly adsorbed by metal-chelating affinity immobilization that is a simple technique with the advantage that support may be reused [54]. Desorption agents may cause the desorption of the enzyme that can be also achieved by changing the pH value [115]. The compound n-hexane could reduce the deactivation of AGMNP-CO2 <sup>+</sup>-PFL from methanol. It was reported that biodiesel production from oil transesterification was higher with n-hexane than using tert-butanol [70,94].

Encapsulation immobilization entraps lipase by a co-precipitation method and crosslinking agents like glutaraldehyde are used to interconnect the enzymes. However, there is a high mass transfer resistance that reduces its efficiency.

The entrapment of cells having lipase activity appears to be a simple and attractive technique. Lipase immobilized in silicon granules or calcium alginate beads, with glucose supplementation for cells maintenance, achieved an increased number of cycles, 28 instead of 23, while keeping 90% activity [33]. Whole cell, recombinant methods and protein and metabolic engineering are promising options to increase lipase applications [82].

### *5.2. Magnetic Nanocarriers*

Materials like magnetite (Fe3O4) are used as support for immobilization because they allow a rapid separation from the reaction medium when an external magnetic field is applied [116]. The development of magnetic nanoparticles (MNPs) as a support for enzymes immobilization has been recently reviewed [117]. Typical magnetic nanomaterials include iron oxide (Fe3O4 and γ-Fe2O3), alloy-based (CoPt3 and FePt), pure metal (Fe and Co), and spinel-type ferromagnet (MgFe2O4, MnFe2O4, and CoFe2O4) [117]. Examples of lipase immobilized in various types of magnetic nanoparticles are shown in Table 4.


**Table 4.** Biodiesel production with various immobilized lipases using magnetic nanoparticles as carriers.

<sup>a</sup> Biodiesel yield (wt/wt.%) was determined as the methyl esters amount produced by the lipases in the reaction process divided by the initial amount of esters.

Magnetic nanoparticles (MNPs) have good biocompatibility and non-toxicity but tend to aggregate and oxidize, so they need to be functionalized on the surface and use a cross-linking agent to bind the enzyme. One way is by using silica coating where a silica shell is formed on the surface by using amino-functional reagents like 3-aminopropyl triethoxysilane (APTES). Fe3O4 particles were encapsulated with mesoporous silicon and modified with APTES or 3-mercaptopropyl trimethoxysilane (MPTMS) followed by binding of the lipase with glutaraldehyde. APTES-modified Fe3O4 particles were reported to give better yield of biodiesel (90%) than MPTMS particles [75]. *Burkholderria* sp. lipase on Fe3O4 MNPs also achieved 90% conversion [123].

Another way to protect MNPs is with organic polymers, including synthetic polymers and biopolymers. The polymer can be either incorporated into the precursor solution to form Fe3O4 MNPs or externally to create the core shell [117]. The polymer surface provides numerous functional groups that facilitate the enzyme binding. Magnetic chitosan binds the lipase by covalent attachment [124].

Separation of nanobiocatalysts is difficult in an oily system [119] but the magnetic properties of MNPs can facilitate the separation of enzyme from reaction media. In this way, the reaction may be immediately finished as well as using the enzyme for further uses [117].

### *5.3. Coimmobilization*

Some works have proposed to use coimmobilization of enzymes. The advantage is that the first enzyme releases the product that is transferred to the next coimmobilized enzyme with a short diffusional distance. This is especially relevant for lipases due to their specificity for triacylglycerols ester bonds. The mixture of 1,3-specific lipase and a non-specific lipase enhances the global activity because it removes the limiting acyl-migration step. Several coimmobilized systems have been studied for biodiesel production like *Candida rugosa* lipase and *Rhyzopus oryzae* lipase simultaneously on silica gel [74], *Candida antarctica* lipase B and *Thermomyces lanuginosus* lipase, on the surface of the *Phichia pastoris* cell [57], and *Rhizomucor miehei* and *Candida antarctica* lipases on epoxy-functionalized silica [67,100].

Lipases are coimmobilized on the same material surface in order to get better global activity and improved enzyme specificity and selectivity for hydrolysis of triacylglycerols as well as those generated diacylglycerols and monoacylglycerols that must be further hydrolyzed [96]. However, the most active enzyme may get some loss of activity through this procedure [125]. A different coimmobilization strategy was proposed by immobilizing several lipases layer-by-layer using abcoating with polyethylenimine [126]. Other authors have used coating with PEI/glutaraldehyde to form 5 enzyme layers of lipases A and B from *Candida antarctica*, lipases from *Rhizomocur miehei*, and *Themomyces lanuginosus* and phospholipase Lecitase Ultra [125]. Although it gives an innovative way for fats hydrolysis some problems might arise either from inhibition by coating agents used, high costs of different lipases used or steric hindrance for accessibility of triacylglycerols to the active site of lipases immobilized in the inner layers.

### **6. Industrial Applications of Lipase-Catalyzed Biodiesel**

Even though transesterification through alkaline catalysis is the preferred process in the majority of industrial biodiesel production plants [23], a few lipase-based processes have already been implemented to plant-scale operation. The collaboration of Novozymes (Bagsvaerd, Denmark) with Piedmont Biofuels (Pittsboro, NC, USA) resulted in a patent application to produce fatty acid alkyl esters, by a lipolytic enzyme in a solution containing triacylglycerol, alcohol, water, and glycerol [93,127]. Viesel Fuel (Terrac Stuart, FL, USA) upgraded in 2013 its facility through an enzymatic process developed by Novozymes (Denmark) to use brown grease and waste cooking oil to produce up to 11 million gallons biodiesel per year using Eversa Transform® lipase from Novozymes, a soluble lipase produced by a genetically modified strain of *Aspergillus oryzae* [128], and an ion exchange resin system for removal of remaining free fatty acids during crude biodiesel refining [129,130]. Viesel Fuel, Novozymes and Tactical Fabrication also collaborated with Buster Biofuels to upgrade its facility in San Diego (CA, USA) to produce up to 5 million gallons per year [131]. Lvming and Environmental Protection Technology Co. Ltd. (Shanghai, China) used lipase of *Candida* sp. to produce 10,000 tons per year from waste frying oil [95]. A plant in Sumaré (Sao Paulo, Brazil) produces biodiesel from mixed beef tallow and soybean oil using Callera® Trans L lipase in a batch reactor [132]. These companies are using liquid lipase formulations but the efficiency of the process can be improved further by using recent developments in immobilized lipases. So, Hunan Rivers Bioengineering Co. Ltd. (Hunan, China) was reported to use Novozym 435® lipase in a stirred tank reactor to produce 20,000 tons of biodiesel per year. The enzyme is a lipase B from *Candida antarctica* immobilized on a resin consisting of macroporous support formed by poly(methyl methacrylate) crosslinked with divinylbenzene [133]. New technology protected with patents [134] has been provided by EnzymoCore, a leading global

producer company founded in 2007 in Israel and with several active biodiesel plants around the world. This company has developed modified-immobilized enzymes, supported on solid organic resins, with high resistance to methanol and able to produce biodiesel from any type of oil or fat, even those cheap and with very large content of free fatty acids and polar lipids [135].

### **7. Conclusions**

Animal fat waste, usually resulting from slaughterhouses, the meat processing industry, and cooking facilities, is being increasingly used for biodiesel production. Transesterification through alkaline catalysis is the preferred process at industrial biodiesel production plants. Transesterification with lipases has not been attractive for industry yet because of the higher operative costs in comparison to alkaline catalysis; transesterification with lipases has problems including poor enzyme stability, difficulties in reusability, and denaturation by alcohol although they are not affected by water and free fatty acids typically found in animal fats. However, a few companies could solve such troubles since they are running liquid lipase formulations for producing biodiesel from cooking oil waste at industrial scale. However, the efficiency of the process can be further improved. Recent developments in immobilized lipases and availability of different types of supports such as mesoporus materials, silica nanoflowers, pickering emulsion, and metal-organic frameworks demonstrate improved efficiency and reduced costs. Immobilization of the enzyme in such materials increases its stability and makes it more resistant to denaturation by alcohol. Magnetic nanomaterials constitute an even better support for enzyme immobilization because they can be recovered when an external magnetic field is applied. These nanoparticles are functionalized on the surface by coating with silica or organic polymers that enhance the efficiency of the process. The entrapment of whole cells with lipase activity, appears to be simple and efficient although more research is needed. Coimmobilization of lipases is an innovative process, but not so attractive for industrial application. It needs further research because of the need for different lipases that increases the costs and the efficiency affected by steric difficulties for enzymes to hydrolyze triacylglycerols.

**Author Contributions:** Conceptualization, F.T.; resources, F.T.-R. and L.M.; writing—original draft preparation, F.T.-R.; writing—review and editing, F.T.-R., L.M. and F.T.; supervision, F.T. All authors have read and agreed to the published version of the manuscript.

**Funding:** This research was funded by European Marie Curie project, grant number 614281 (HIGHVALFOOD) and European Regional Development Fund.

**Conflicts of Interest:** The authors declare no conflict of interest. The funders had no role in the design of the study; they had no role in the collection, analyses, or interpretation of data; and they had no role in the writing of the manuscript, or in the decision to publish the results.

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